*>EPA
United States
Environmental
Protection Agency
EPA/600/R-23/375 | January 2024 | www.epa.gov/isa
Pb
Lead
Office of Research and Development
Center for Public Health & Environmental Assessment
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United States
Environmental Protection
Agency
EPA/600/R-23/375
January 2024
www.epa.gov/isa
Integrated Science
Assessment for Lead
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
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DISCLAIMER
This document has been reviewed in accordance with the U.S. Environmental Protection Agency
policy and approved for publication. Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
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DOCUMENT GUIDE
This Document Guide is intended to orient readers to the organization of the Lead (Pb) Integrated
Science Assessment (ISA) in its entirety and to the sub-section of the ISA at hand (indicated in bold). The
ISA consists of the Front Matter (list of authors, contributors, reviewers, and acronyms), Executive
Summary, Integrated Synthesis, and 12 appendices, which can all be found at https://assessments.epa.gov/
isa/document/&deid=3 59536.
Front Matter
Executive Summary
Integrated Synthesis
Appendix 1. Lead Source to Concentration
Appendix 2. Exposure, Toxicokinetics, and Biomarkers
Appendix 3. Nervous System Effects
Appendix 4. Cardiovascular Effects
Appendix 5. Renal Effects
Appendix 6. Immune System Effects
Appendix 7. Hematological Effects
Appendix 8. Reproductive and Developmental Effects
Appendix 9. Effects on Other Organ Systems and Mortality
Appendix 10. Cancer
Appendix 11. Effects of Lead in Terrestrial and Aquatic Ecosystems
Appendix 12. Process for Developing the Pb Integrated Science Assessment
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CONTENTS
DOCUMENT GUIDE Mi
LIST OF TABLES vi
LIST OF FIGURES viii
INTEGRATED SCIENCE ASSESSMENT TEAM FOR LEAD ix
AUTHORS, CONTRIBUTORS, AND REVIEWERS xii
ACRONYMS AND ABBREVIATIONS xix
EXECUTIVE SUMMARY ES-1
ES.1 Purpose and Scope of the Integrated Science Assessment for Lead ES-1
ES.2 Pb in Ambient Air ES-2
ES.3 Fate and Transport ES-3
ES.4 Trends ES-4
ES.5 Exposure ES-5
ES.6 Health and Welfare Effects of Pb Exposure ES-7
ES.6.1 Health Effects of Pb Exposure ES-7
ES.6.2 Welfare Effects of Pb Exposure ES-11
ES.7 Key Aspects of Health and Welfare Effects Evidence ES-16
ES.7.1 Health Effects Evidence: Key Findings ES-16
ES.7.2 Welfare Effects Evidence: Key Findings ES-18
ES.8 References ES-20
INTEGRATED SYNTHESIS FOR LEAD IS-1
15.1 Introduction IS-2
15.1.1 Purpose and Overview IS-2
15.1.2 Pb Integrated Science Assessment Process and Development IS-3
15.2 Pb Source to Concentration IS-8
15.2.1 Sources and Emissions IS-9
15.2.2 Fate and Transport IS-9
15.2.3 Sampling and Analysis IS-11
15.2.4 Ambient Air Pb Concentrations IS-12
15.3 Trends IS-13
15.4 Human Exposure to Ambient Pb IS-17
15.5 Toxicokinetics IS-18
15.6 Pb Biomarkers IS-20
15.7 Evaluation of the Health Effects of Pb IS-21
15.7.1 Connections Among Health Effects IS-21
15.7.2 Biological Plausibility IS-22
15.7.3 Summary of Health Effects Evidence IS-24
15.7.4 At-Risk Populations IS-62
15.8 Evaluation of Welfare Effects of Pb IS-74
15.8.1 Summary of Effects on Terrestrial Ecosystems IS-76
15.8.2 Summary of Effects on Freshwater Ecosystems IS-78
15.8.3 Summary of Effects on Saltwater Ecosystems IS-81
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IS.8.4 Summary of Welfare Effects Evidence IS-84
IS.9 Policy-Relevant Issues IS-101
15.9.1 Air Pb-to-Blood Pb Relationships IS-102
15.9.2 Concentration-Response Relationships for Human Health Effects IS-104
15.9.3 Lifestages and Timing of Pb Exposure Contributing to Observed Nervous System
Effects IS-106
15.9.4 Ecological Effects and Corresponding Pb Concentrations IS-107
IS. 10 References IS-108
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LIST OF TABLES
Table IS-1
Table IS-2A
Table IS-2B
Table IS-2C
Table IS-2D
Table IS-2E
Summary of causality determinations by health outcome_
IS-25
Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and nervous system effects ascertained during childhood, adolescent, and young adult
lifestages IS-28
Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and nervous system effects ascertained during childhood, adolescent, and young adult
lifestages IS-30
Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and nervous system effects ascertained during childhood, adolescent, and young adult
lifestages IS-32
Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and nervous system effects ascertained during childhood, adolescent, and young adult
lifestages IS-33
Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and nervous system effects ascertained during childhood, adolescent, and young adult
lifestages IS-34
Table IS-2F Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and nervous system effects ascertained during childhood, adolescent, and young adult
lifestages IS-36
Table IS-3A Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and nervous system effects ascertained during adult lifestages IS-39
Table IS-3B Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and nervous system effects ascertained during adult lifestages IS-40
Table IS-4 Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and cardiovascular effects and cardiovascular-related mortality IS-43
Table IS-5 Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and renal effects IS-44
Table IS-6A Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and immune system effects IS-46
Table IS-6B Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and immune system effects IS-48
Table IS-7 Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and hematological effects IS-50
Table IS-8A Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and reproductive and developmental effects IS-52
Table IS-8B Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and reproductive and developmental effects IS-54
Table IS-8C Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and reproductive and developmental effects IS-55
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Table IS-8D Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and reproductive and developmental effects IS-57
Table IS-9 Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and musculoskeletal effects IS-58
Table IS-10 Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and total (nonaccidental) mortality IS-60
Table IS-11 Summary of evidence from epidemiologic and animal toxicological studies on Pb exposure
and cancer IS-61
Table IS-12
Table IS-13
Table IS-14
Table IS-15
Table IS-16
Table IS-17
Table IS-18
Table IS-19
Table IS-20
Table IS-21
Characterization of evidence for factors potentially increasing the risk for Pb-related health
effects IS-63
Summary of evidence for population characteristics and other factors potentially related to
increased risk of Pb-related health effects IS-66
Summary of causality determinations for welfare effects of Pb
Summary of evidence for effects of Pb on hematological endpoints in terrestrial and
aquatic biota
Summary of evidence for effects of Pb on neurobehavioral endpoints in terrestrial and
aquatic biota
IS-85
Summary of evidence for effects of Pb on physiological stress endpoints in terrestrial and
aquatic biota IS-86
IS-88
IS-90
Summary of evidence for effects of Pb on survival of terrestrial and aquatic biota IS-92
Summary of evidence for growth effects of Pb in terrestrial and aquatic biota IS-95
Summary of evidence for reproductive effects of Pb in terrestrial and aquatic biota IS-98
Summary of evidence for community and ecosystem effects of Pb IS-101
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LIST OF FIGURES
Figure ES-1 Conceptual model of air-related Pb exposure through inhalation and ingestion. ES-6
Figure ES-2 Summary of causality determinations by health outcome. ES-8
Figure ES-3 Summary of causality determinations for ecological effects of Pb. ES-12
Figure IS-1 Pb maximum rolling 3-month average in |jg/m3 for the 2020-2022 period. IS-12
Figure IS-2 Maps of Pb sampled from (A) A-horizon and (B) C-horizon soils, (C) the ratio of Pb
observed in A-horizon to C-horizon soils, and (D) population density. IS-15
Figure IS-3 Illustrative figure for potential biological pathways for health effects following Pb exposure. IS-23
Figure IS-4 Slope factors for blood Pb as a function of air Pb. IS-103
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INTEGRATED SCIENCE ASSESSMENT TEAM FOR LEAD
Executive Direction
Dr. Steven J. Dutton (Director)—Health and Environmental Effects Assessment Division,
Center for Public Health and Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Christopher P. Weaver (Director)—Integrated Climate Sciences Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. John Vandenberg (Director)—former Health and Environmental Effects Assessment
Division, Center for Public Health and Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Emily Gibb Snyder (Associate Director)—Health and Environmental Effects
Assessment Division, Center for Public Health and Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research Triangle
Park, NC
Dr. Tara Greaver (Acting Branch Chief)—Integrated Climate Sciences Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Scott Jenkins (Branch Chief)—Health and Environmental Effects Assessment Division,
Center for Public Health and Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Jane Ellen Simmons (Branch Chief)—former Health and Environmental Effects
Assessment Division, Center for Public Health and Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research Triangle
Park, NC
Technical Support Staff
Ms. Christine Alvarez—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Andrea Bartolotti—Research Planning and Implementation Staff, Center for
Environmental Measurement and Modeling, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Olivia Birkel—Oak Ridge Associated Universities, Health and Environmental Effects
Assessment Division, Center for Public Health and Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Washington, DC
Ms. Marieka Boyd—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. LaShonda Bunch—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
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Ms. Shannon Cassell—former Oak Ridge Associated Universities, Health and
Environmental Effects Assessment Division, Center for Public Health and Environmental
Assessment, Office of Research and Development, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Ms. Kate Conrad—Oak Ridge Associated Universities, Health and Environmental Effects
Assessment Division, Center for Public Health and Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research Triangle
Park, NC
Ms. Megan Dunne—former Oak Ridge Associated Universities, Health and Environmental
Effects Assessment Division, Center for Public Health and Environmental Assessment,
Office of Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Ms. Amanda Haddock—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Cheryl Itkin—Research Planning and Implementation Staff, Center for Public Health
and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Washington, DC
Ms. Maureen Johnson—Research Assessment and Communications Outreach Staff, Center
for Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Washington, DC
Mr. Ryan Jones—Health and Environmental Effects Assessment Division, Center for Public
Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Regina Madalena—Research Planning and Implementation Staff, Center for Public
Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Danielle Moore—former Senior Environmental Employment Program, Health and
Environmental Effects Assessment Division, Center for Public Health and Environmental
Assessment, Office of Research and Development, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Mr. Sam Penry—Oak Ridge Associated Universities, Integrated Climate Sciences Division,
Center for Public Health and Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Mira Sanderson—Oak Ridge Associated Universities, Health and Environmental
Effects Assessment Division, Center for Public Health and Environmental Assessment,
Office of Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Ms. Karlee Shadle—Office of Science Advisor, Policy and Engagement, Office of Research
and Development, U.S. Environmental Protection Agency, Washington, DC
Ms. Kathleen Shank—Oak Ridge Associated Universities, Integrated Climate Sciences
Division, Center for Public Health and Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Jenna Strawbridge—former Oak Ridge Associated Universities, Health and
Environmental Effects Assessment Division, Center for Public Health and Environmental
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Assessment, Office of Research and Development, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Ms. Jessica Taylor Anders—Integrated Climate Sciences Division, Center for Public Health
and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Washington, DC
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AUTHORS, CONTRIBUTORS, AND REVIEWERS
Authors
Mr. Evan Coffman1' (Health Assessment Team Lead, Integrated Science Assessment for
Lead)—Health and Environmental Effects Assessment Division, Center for Public Health
and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Meredith Lass iter' (Welfare Assessment Team Lead, Integrated Science Assessment for
Lead)—Integrated Climate Sciences Division, Center for Public Health and
Environmental Assessment, Office of Research and Development, U.S. Environmental
Protection Agency, Research Triangle Park, NC
Ms. Anna Champlin1' (Project Manager, Integrated Science Assessment for Lead)—Health
and Environmental Effects Assessment Division, Center for Public Health and
Environmental Assessment, Office of Research and Development, U.S. Environmental
Protection Agency, Washington, DC
Dr. Timothy Anderson—Health and Environmental Impacts Division, Office of Air Quality
Planning and Standards, Office of Air and Radiation, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Mr. Michael Beuthe—Technology Innovation and Field Services Division, Office of
Superfund Remediation and Technology Innovation, Office of Land and Emergency
Management, U.S. Environmental Protection Agency, Edison, NJ
Ms. Katie Boaggio— Sector Policies and Programs Division, Office of Air Quality Planning
and Standards, Office of Air and Radiation, U.S. Environmental Protection Agency,
Research Triangle Park, NC
Dr. James Brown—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Barbara Buckley—former Health and Environmental Effects Assessment Division,
Center for Public Health and Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Peter Byrlcy '—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Laura M. Carlson—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Catheryne Chiang—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
t Appendix Lead
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Dr. Rebecca Dalton—Integrated Climate Sciences Division, Center for Public Health and
Environmental Assessment, Office of Research and Development, U.S. Environmental
Protection Agency, Research Triangle Park, NC
Dr. Stephanie DcFlorio-Barkcr'—Public Health and Environmental Systems Division,
Center for Public Health and Environmental Assessment, Office of Research and
Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Jean-Jacques Dubois—Health and Environmental Effects Assessment Division, Center
for Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Parker F. Duffney—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Zahra Gohari—former Oak Ridge Associated Universities, Health and Environmental
Effects Assessment Division, Center for Public Health and Environmental Assessment,
Office of Research and Development, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Ms. Rebecca Gray—ICF, Durham, NC
Dr. Brooke L. Hemming—Integrated Climate Sciences Division, Center for Public Health
and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Kirstin Hester—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Anthony Jones—Health and Environmental Impacts Division, Office of Air Quality
Planning and Standards, Office of Air and Radiation, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Dr. S. Douglas Kaylor—Integrated Climate Sciences Division, Center for Public Health and
Environmental Assessment, Office of Research and Development, U.S. Environmental
Protection Agency, Research Triangle Park, NC
Ms. Haesoo Kim—Oak Ridge Associated Universities, Health and Environmental Effects
Assessment Division, Center for Public Health and Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research Triangle
Park, NC
Dr. Ellen Kirranc'—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Alison Krajewski'—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Nichole Kulikowski—Health and Environmental Effects Assessment Division, Center
for Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Washington, DC
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Dr. Archana Lamichhane—Health and Environmental Impacts Division, Office of Air
Quality Planning and Standards, Office of Air and Radiation, U.S. Environmental
Protection Agency, Research Triangle Park, NC
Ms. Emma Leath—Oak Ridge Associated Universities, Integrated Climate Sciences
Division, Center for Public Health and Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. David M. Lchmann '—Health and Environmental Effects Assessment Division, Center
for Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Cynthia Lin—ICF, Durham, NC
Dr. Qingyu Meng—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Stephen McDow'—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Leigh C. Moorhead—Integrated Climate Sciences Division, Center for Public Health
and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Anuradha Mudipalli—Health and Environmental Effects Assessment Division, Center
for Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Natalia Neal-Walthall—former Oak Ridge Associated Universities, Health and
Environmental Effects Assessment Division, Center for Public Health and Environmental
Assessment, Office of Research and Development, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Dr. Kristopher Novak—Integrated Climate Sciences Division, Center for Public Health and
Environmental Assessment, Office of Research and Development, U.S. Environmental
Protection Agency, Research Triangle Park, NC
Dr. Nicole Olson—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Russell D. Owen—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Michael Pennino—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Washington, DC
Mr. R. Byron Rice—Integrated Climate Sciences Division, Center for Public Health and
Environmental Assessment, Office of Research and Development, U.S. Environmental
Protection Agency, Research Triangle Park, NC
Dr. Rachel M. Shaffer—Chemical Pollutant Assessment Division, Center for Public Health
and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Washington, DC
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Dr. Michael Stew art '—Research Planning & Implementation Staff, Center for Public Health
and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Contributors
Ms. Meredith Clemons—ICF, Durham, NC
Ms. Tamara Dawson—ICF, Durham, NC
Dr. Sorina Eftim—ICF, Durham, NC
Ms. Julia Finver—ICF, Durham, NC
Ms. Alexandra Goldstone—ICF, Durham, NC
Ms. Tara Hamilton—ICF, Durham, NC
Mr. Anthony Hannani—ICF, Durham, NC
Mr. Max Hatala—Oak Ridge Associated Universities, Health and Environmental Effects
Assessment Division, Center for Public Health and Environmental Assessment, Office of
Research and Development, U.S. Environmental Protection Agency, Research Triangle
Park, NC
Ms. Michele Justice—ICF, Durham, NC
Ms. Ila Kanneboyina—Oak Ridge Associated Universities, Health and Environmental
Effects Assessment Division, Center for Public Health and Environmental Assessment,
Office of Research and Development, U.S. Environmental Protection Agency,
Washington, DC
Ms. Afroditi Kastigiannakis—ICF, Durham, NC
Ms. Anna Kolanowski—ICF, Durham, NC
Ms. Madison Lee—ICF, Durham, NC
Dr. Nathan Lothrop—ICF, Durham, NC
Ms. Denyse Marquez Sanchez—ICF, Durham, NC
Dr. Michelle Mendez—ICF, Durham, NC
Ms. Melissa Miller—ICF, Durham, NC
Ms. Danielle Moore—ICF, Durham, NC
Mr. Kevin O'Donovan—ICF, Durham, NC
Ms. Emily Pak—ICF, Durham, NC
Ms. Jennifer Powers—ICF, Durham, NC
Ms. Sheerin Shirajan—ICF, Durham, NC
Ms. Swati Sriram—ICF, Durham, NC
Ms. Nkoli Ukpabi—ICF, Durham, NC
Dr. Janielle Vidal—ICF, Durham, NC
Ms. Leah West—ICF, Durham, NC
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Ms. Connie Xiong—ICF, Durham, NC
Ms. Maricruz Zarco—ICF, Durham, NC
Reviewers
Mr. Colin Barrette—Air Quality Assessment Division, Office of Air Quality Planning and
Standards, Office of Air and Radiation, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Britta Bierwagen (Associate Director)—Integrated Climate Sciences Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Washington, DC
Dr. Joseph Braun—Brown University
Dr. Kevin Brix—University of Miami
Mr. Halil Cakir—Air Quality Assessment Division, Office of Air Quality Planning and
Standards, Office of Air and Radiation, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Mr. Kevin Cavender—Air Quality Assessment Division, Office of Air Quality Planning and
Standards, Office of Air and Radiation, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Jasim Chowdhury—International Lead Association
Ms. Rebecca Daniels—Public Health and Integrated Toxicology Division, Center for Public
Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Cliff Davidson—Syracuse University
Dr. Rodney Dietert—Cornell University College of Veterinary Medicine
Dr. Cassandra Elizalde—Health and Environmental Impacts Division, Office of Air Quality
Planning and Standards, Office of Air and Radiation, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Dr. Aimen Farraj—Public Health and Integrated Toxicology Division, Center for Public
Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Gabriel Filippelli—Indiana University-Purdue University Indianapolis School of Science
Ms. Jessica Frank—Science Policy Division, Office of Science Advisor, Policy and
Engagement, U.S. Environmental Protection Agency, Washington, DC
Dr. Eliseo Guallar—Johns Hopkins University
Dr. Iman Hassan—Health and Environmental Impacts Division, Office of Air Quality
Planning and Standards, Office of Air and Radiation, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Ms. Beth Hassett-Sipple—former Integrated Climate Sciences Division, Center for Public Health
and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
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Dr. Erin Hines—Public Health and Integrated Toxicology Division, Center for Public Health
and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Andrew Hotchkiss—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Marion Hoyer—Assessment and Standards Division, Office of Transportation and Air
Quality, Office of Air and Radiation, U.S. Environmental Protection Agency, Ann Arbor,
MI
Dr. Mary Hutson—Health and Environmental Impacts Division, Office of Air Quality
Planning and Standards, Office of Air and Radiation, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Dr. Anthony Jones—Health and Environmental Impacts Division, Office of Air Quality
Planning and Standards, Office of Air and Radiation, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Ms. Hali Kerr, Attorney-Advisor—Air and Radiation Law Office, Office of General
Counsel, U.S. Environmental Protection Agency, Washington, DC
Dr. Roman Lanno—Ohio State University
Dr. Qingyu Meng*—Health and Environmental Effects Assessment Division, Center for
Public Health and Environmental Assessment, Office of Research and Development,
U.S. Environmental Protection Agency, Research Triangle Park, NC
Ms. Leigh Meyer—Health and Environmental Impacts Division, Office of Air Quality
Planning and Standards, Office of Air and Radiation, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Dr. Deirdre Murphy—Health and Environmental Impacts Division, Office of Air Quality
Planning and Standards, Office of Air and Radiation, U.S. Environmental Protection
Agency, Research Triangle Park, NC
Dr. Liz Naess (Acting Branch Chief)—Health and Environmental Effects Assessment
Division, Center for Public Health and Environmental Assessment, Office of Research
and Development, U.S. Environmental Protection Agency, Research Triangle Park, NC
Dr. Zachary Pekar—Environmental Impacts Division, Office of Air Quality Planning and
Standards, Office of Air and Radiation, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Mr. Venkatesh Rao—Air Quality Assessment Division, Office of Air Quality Planning and
Standards, Office of Air and Radiation, U.S. Environmental Protection Agency, Research
Triangle Park, NC
Dr. Justin Richardson—University of Massachusetts Amherst
Dr. Jennifer Richmond-Bryant—North Carolina State University
Mr. Dahnish Shams (Acting Associate Director)—Health and Environmental Effects
Assessment Division, Center for Public Health and Environmental Assessment, Office of
Dr. Meng was affiliated with CalEPA at the time of the exposure topic of the Peer Input Workshop that took place
on June 29, 2022.
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Research and Development, U.S. Environmental Protection Agency, Research Triangle
Park, NC
Dr. Christina Sobin—University of Texas at El Paso
Dr. Aaron Specht—Purdue University
Dr. Jay Turner—Washington University - McKelvey School of Engineering
Dr. Rosalind Wright—Icahn School of Medicine at Mount Sinai
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ACRONYMS AND ABBREVIATIONS
AD Alzheimer's disease
ADHD attention-deficit/hyperactivity disorder
AL allostatic load
ALAD S-aminolevulinic acid dehydratase
APOE apolipoprotein E
AQCD Air Quality Criteria Document
AQS Air Quality System
As arsenic
AWQC ambient water quality criteria
BAEP brainstem auditory evoked potential
BLL blood lead level
BLM biotic ligand model
BMD benchmark dose
BMI body mass index
BMS Baltimore Memory Study
BP blood pressure
Ca2+ calcium ion(s)
CAD coronary artery disease
CASAC Clean Air Scientific Advisory
Committee
Cd cadmium
CEC cation exchange capacity
CHD coronary heart disease
CI confidence interval
C-R concentration-response
CVD cardiovascular disease
d day(s)
DOC dissolved organic carbon
DTH delayed-type hypersensitivity
EC 10 effect concentration at 10% inhibition
EC50 half maximal effect concentration
Fe iron
FRM Federal Reference Method
FSIQ full-scale intelligence quotient
GABA gamma-aminobutyric acid
GRIN glutamate ionotropic receptor N-methyl
D aspartate-type subunit
GST glutathione S-transferase
HAZ height-for-age Z-score
Hb hemoglobin
HERO Health and Environmental Research
Online
HFE hemochromatosis gene
Hg mercury
HISA Highly Influential Scientific
Assessment
HMOX1 heme oxygenase-1
hr hour(s)
HRV heart rate variability
IEUBK Integrated Exposure Uptake Biokinetic
IFN-y interferon-gamma
Ig immunoglobulin
IHD ischemic heart disease
IL-4 interleukin-4
IQ intelligence quotient
IRP Integrated Review Plan
IS Integrated Synthesis
ISA Integrated Science Assessment
KNHANES Korea National Health and Nutrition
Examination Survey
LC lethal concentration
LECES Level of Biological Organization,
Exposure, Comparison, Endpoint, and
Study Design
LOEC lowest observed effect concentration
MDI Mental Development Index
Mg2+ magnesium ion
MI myocardial infarction
mo month(s)
Mn manganese
mtDNA mitochondrial DNA
NAAQS National Ambient Air Quality
Standards
NAS Normative Aging Study
NASCAR National Association for Stock Car
Auto Racing
NASGLP North American Soil Geochemical
Landscapes Project
NEI National Emissions Inventory
NHANES National Health and Nutrition
Examination Survey
NOAA National Oceanic and Atmospheric
Administration
NOEC no-observed-effect concentration
OM organic matter
OMB Office of Management and Budget
Pb lead
PDI Psychomotor Developmental Index
PECOS Population, Exposure, Comparison,
Outcome, and Study Design
PHQ Patient Health Questionnaire
PM particulate matter
PP pulse pressure
xix
-------
PQAPP
Program-level QA Project Plan
Th
T helper
QA
quality assurance
TSP
total suspended particulate
RBC
red blood cell
IT
tetanus toxoid
SE
standard error
U.S. EPA
United States Environmental Protection
SES
socioeconomic status
Agency
SHEDS
Stochastic Human Exposure and Dose
VDR
vitamin D receptor
Simulation
wk
week(s)
SNP
single nucleotide polymorphism
yr
year(s)
XX
-------
EXECUTIVE SUMMARY
ES.1 Purpose and Scope of the Integrated Science Assessment for
Lead
The Federal Clean Air Act requires the United States Environmental Protection Agency
(U.S. EPA) to set National Ambient Air Quality Standards (NAAQS) for criteria air pollutants that are
considered harmful to public health and the environment, including lead (Pb), and to periodically review
the science upon which the NAAQS are based. Pb emitted into air can be inhaled or ingested or can
deposit and accumulate in other environmental media (e.g., soil, water, sediment, biota), contributing to a
wide range of effects in humans and wildlife. This Integrated Science Assessment (ISA), prepared by the
U.S. EPA, is a synthesis and evaluation of the most policy-relevant science that forms the scientific
foundation for the review of the primary (health-based) and secondary (welfare-based) NAAQS for Pb.
The Pb primary NAAQS is established to protect public health, including at-risk populations, with an
adequate margin of safety. The Pb secondary NAAQS is intended to protect the public welfare from
known or anticipated adverse effects of Pb in the ambient air and, in this regard, this ISA focuses
specifically on ecological effects.
This Executive Summary (ES) provides an overview of the important conclusions drawn in the
ISA across scientific disciplines, beginning with information on sources of Pb emissions in ambient air,
the fate and transport of Pb in the environment, concentration trends of Pb in air and non-air media, and
pathways of exposure, followed by the health and welfare effects of Pb. Health effects evidence evaluated
in the Pb ISA includes experimental animal studies and observational epidemiologic studies. Welfare
effects evidence evaluated in the Pb ISA includes experimental and observational studies examining the
effects of Pb on terrestrial, freshwater, and saltwater ecosystems and biota. A more extensive summary of
the evidence and conclusions of the Pb ISA is presented in the Integrated Synthesis (IS), and detailed
study-level information and an in-depth characterization of the weight-of-evidence conclusions are
included in individual appendices for each topic area. Studies considered in the development of the ISAs
are documented in the U.S. EPA Health and Environmental Research Online (HERO) database. The
publicly accessible HERO project page for this ISA contains the references that were considered for
inclusion and provides bibliographic information and abstracts.
The previous ISA for Pb was published in 2013 (U.S. EPA. 2013) and included peer-reviewed
literature published through September 2011. Prior Pb assessments include the 2006 Air Quality Criteria
Document (AQCD) for Pb (U.S. EPA. 2006). the 1986 Pb AQCD (U.S. EPA. 1986a) and its associated
addendum (U.S. EPA. 1986b). the 1990 Supplement to the 1986 addendum (U.S. EPA. 1990). and the
1977 Pb AQCD (U.S. EPA. 1977). The most recent review of the primary and secondary Pb NAAQS was
completed in 2016, at which time the existing standards from 2008 were retained without revision (81 FR
71906). In the 2008 review, the interpretation of the science in the 2006 Pb AQCD led to the decision to
ES-1
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lower the levels of the primary and secondary NAAQS for Pb by ten-fold, from the 1978 levels of
1.5 (ig/m3 to a level of 0.15 |ig/nr\ The averaging time was revised to a rolling three-month period with a
maximum (not-to-be-exceeded) form, evaluated over a three-year period. U.S. EPA's decision to revise
the primary standard in 2008 was based on the substantially expanded body of health effects evidence
available at that time, including evidence for cognitive effects of Pb exposure in children. The revised
2008 standard was established to increase protection against air Pb related human health effects, including
neurocognitive effects, for children and other at-risk populations. In 2016, the U.S. EPA Administrator
concluded that the existing primary standard provides health protection from air emissions for Pb for at-
risk groups, especially children, and the existing secondary standard provides protection against adverse
effects to public welfare from air emissions for Pb, including harm to aquatic and terrestrial ecosystems
(81 FR 71906).
This ISA focuses on synthesizing and integrating the evidence that has become available since the
2013 Pb ISA with the information and conclusions from previous assessments. Key policy-relevant
conclusions are intended to inform the U.S. EPA's review of the Pb NAAQS, including conclusions on
the populations at increased risk of Pb-related effects, the Pb exposure concentrations at which such
effects occur, and the overall strength of the evidence supporting relationships between Pb exposures and
health or welfare effects. Conclusions on the overall strength of evidence are described using a five-level
hierarchy that classifies the weight of evidence for causation into one of the following categories:
• Causal relationship
• Likely to be a causal relationship
• Suggestive of, but not sufficient to infer, a causal relationship
• Inadequate to infer a causal relationship
• Not likely to be a causal relationship
These causality determinations are made for broad health and welfare effect categories and are
informed by evaluating evidence across scientific disciplines for consistency, coherence, and biological
plausibility, as well as for uncertainties. The ISA's approach to evaluating the weight of evidence and
reaching causality determinations is described in more detail in the Preamble to the Integrated Science
Assessments (U.S. EPA. 2015).
ES.2 Pb in Ambient Air
Exposure to Pb can occur from contaminated air, water, soil, and dust. When it is released from
industrial processes into the air, Pb is mainly emitted into the air in particulate form (IS.2.3). In general,
fine particulate Pb is mostly soluble and removed from the atmosphere by wet deposition, and coarse
particulate Pb is mostly insoluble and removed from the atmosphere by dry deposition. Total Pb
emissions have steadily decreased for decades largely due to the elimination of leaded gasoline used in
ES-2
-------
automobiles before 1996, and, in later years, to reductions in emissions from metals processing sources
(U.S. EPA. 2022b. 2013. 2006). From 1990 to 2020, there was a steep decline in total U.S. Pb emissions,
from about 5 kton/year to less than 1 kton/year. Over this time period, industrial sources have been
replaced by non-road mobile sources as the dominant category of Pb emissions, with emissions from
aircraft that operate on leaded aviation fuel as the largest emissions source in this category (U.S. EPA.
2021). Total estimated national air emissions from the 2020 National Emissions Inventory (NEI) were
621 tons, with 69% from emissions associated with use of leaded aviation gasoline, 18% from industrial
sources, including smelting and metals processing, 9% from fuel combustion, and 3% from wildland
fires. All other sources of Pb air emissions combined were estimated to account for about 2% of total U.S.
Pb emissions in the NEI. Pb emissions from residential wood combustion are not included in the 2020
NEI but can also be a source in areas affected by wood smoke in the winter. In addition to contemporary
Pb emissions into the atmosphere, soil-bound Pb near historical sources can potentially become airborne
under some wind or traffic conditions. This resuspended legacy Pb is also not included in the NEI.
Several recent studies indicated substantial spatial variability in urban ambient air Pb
concentrations influenced by proximity to local sources or industrial activities. Across urban and
neighborhood scales, these variations in ambient air Pb concentrations may not be captured by national
monitoring networks. Seasonal trends were reported in numerous recent studies, but results were mixed,
and no consistent national pattern of seasonality was apparent. Shifts in Pb size distributions since the
1980s from a mass median diameter usually smaller than 2.5 (mi to a mass median diameter between 2.5-
10 |im have been documented in ambient air near roads, near industrial sources, in rural locations, and in
urban locations within the United States and the European Union. No recent studies specifically
investigated background Pb concentrations, but a plausible range of 0.2 to 1 ng/m3 was proposed based on
earlier studies in the 2013 Pb ISA (U.S. EPA. 2013).
ES.3 Fate and Transport
Pb emitted into the atmosphere can be distributed into soil, water, and other media (IS.2.4). The
fate and transport of Pb emitted into the air depends on particle size, which in turn depends largely on the
source emissions. Particle-bound Pb associated with fine particulate matter is transported long distances
and can be found in remote areas, while Pb associated with coarse particulate matter is more likely to
deposit closer to its source. After deposition, resuspension of soil-bound Pb can contribute to airborne
concentrations near major Pb sources. Once deposited in soil, Pb is strongly retained in soil organic
material, and subsequent Pb fate and transport through the soil column is influenced by several
physicochemical factors, such as storage in leaf litter, amount and decomposition rates of organic matter,
composition of organic and inorganic soil constituents, microbial activity, and soil pH. These
physicochemical properties are based on soil forming factors (i.e., climate, organisms, parent material,
relief (shape of the landscape), time, and anthropogenic input). Soils that differ in these factors will
subsequently have different physicochemical properties and different trends in Pb transport. In water,
ES-3
-------
runoff from urban or historically industrial areas contains higher Pb concentrations than non-urban areas.
Recent studies have improved our understanding of soil fate and transport in many areas. These include
the relationships between street characteristics, population density, and land cover with runoff. Recent
research has also expanded on the influence of seasonality and precipitation events on runoff. In addition,
there have been advances in research on transport and sedimentation. While Pb deposition has decreased
in the last half century with the phase-out of leaded gasoline and stricter regulation of some Pb sources,
accumulated Pb-contaminated sediments in areas with a history of industry and urbanization are
vulnerable to resuspension and both down and upstream movement following a disturbance event. For
example, dam removal or other disturbances to water bodies can lead to resuspension and dissolution of
Pb-contaminated sediment that was previously deposited. With the predicted increase in drought
alongside less frequent but more severe precipitation patterns across most of the United States, there may
be a potential for remobilization of legacy Pb.
Additional media besides air, water, and soil play a role in understanding how Pb moves and
changes overtime in the urban environment (IS.2.2). Urban soil, resuspended dust, road dust, and house
dust serve as urban compartments between which Pb can be transported or cycled. High Pb concentrations
are characteristic of urban soil in comparison to other soils and are often related to legacy sources. Studies
in several U.S. cities have explored the high spatial variability of urban soil Pb concentrations, with hot
spots related to income and racial disparities. In recent studies, associations between airborne Pb and
elemental indicators of airborne soil have been observed, suggesting the potential for contaminated soil to
be a source of airborne Pb locally in urban and industrial areas under some circumstances. Resuspension
of urban soil can also be a source of Pb in house dust.
ES.4 Trends
Pb concentrations in ambient air in the United States have decreased since the 1970s, mainly due
to the phase-out of Pb in gasoline (IS.3). For some monitors, there has also been a more recent period of
continued decline in ambient concentrations corresponding to reductions in Pb emissions from local and
regional industrial sources. Based on Pb monitoring network data, the national median of maximum 3-
month average Pb concentrations across monitoring sites declined by 89% from 1990 to 2010 for a mix of
74 source-oriented and non-source-oriented monitors that operated continuously through this period
(IS.3). For a smaller population of 37 monitors with a higher proportion of source-oriented monitors that
operated continuously from 2010 to 2021, the national median of maximum 3-month average Pb
concentrations across monitoring sites decreased by 88% over that period (IS.3). This recent decrease was
driven by the 2008 Pb NAAQS revision and the steepest declines were observed over the period from
2012 to 2015 at mostly source-oriented monitors when emissions from nearby sources were being
reduced to meet the new 2008 Pb NAAQS requirements (IS.3). The declining trend since 2010 is
therefore more representative of a small number of communities near major sources than an urban or
national median. A national trend is more difficult to assess because the number of non-source-oriented
ES-4
-------
monitors is small, and their observed concentrations are close to method detection limits on most days
(IS.3).
Changes in the patterns of Pb emissions over time and between regions of the United States are
also detectable in non-air environmental media and biota (IS.3). Pb may be retained in soils, sediments,
the shells of long-lived bivalves, or trees, where it provides a historical record of Pb deposition over
periods of decreasing Pb emissions, such as the phase-out of Pb from on-road gasoline and reductions in
industrial releases. Overall, evidence from national and regional surveys of Pb in environmental media
and biota reflects a decline in anthropogenic emissions of Pb. However, Pb persists in environmental
media and is still observed in measurable concentrations within biota, particularly near historic and
current sources of Pb pollution. Long-term monitoring of Pb concentration trends in biota (e.g., the
National Oceanic and Atmospheric Administration Mussel Watch program) and soil surveys covering
large spatial extents (e.g., the U.S. Geological Survey North American Soil Geochemical Landscapes
Project) provide essential records of Pb concentrations in the environment observed across decades and
regions. Information on atmospheric Pb concentration trends can be difficult to interpret due to the
influence of other anthropogenic inputs of Pb and heterogeneity associated with natural environments.
Despite reductions in Pb pollution in recent decades, anthropogenic Pb persists in the environment.
ES.5 Exposure
Exposures are considered air-related if they pass through the air compartment at any point prior to
plant, animal, or human contact. For example, air-related Pb exposure may occur through direct inhalation
of air that contains Pb or ingestion of food, water, dust and soil, or other materials that have been
contaminated by Pb originally in ambient air. Non-ambient air-related exposures include those from an
occupation, hand-to-mouth contact with Pb-containing consumer goods, hand-to-mouth contact with dust
or chips of peeling Pb-containing paint, or ingestion of Pb in drinking water conveyed through Pb pipes.
Pb body burden is an aggregation of all of these different exposures. Figure ES-1 depicts the various
pathways that ambient air Pb can take through environmental media to reach human beings (IS.4).
ES-5
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This figure displays air-related exposure pathways of Pb through the environment. Dashed lines represent resuspension of Pb into
the air. Green ovals represent sources of Pb that are not associated with the air compartment but may contribute to Pb along the
exposure pathway. Pb from processing includes Pb that may end up in diet as a result of intentional or inadvertent addition of Pb to
food or food additives such as spices. Other recognized sources of Pb exposure such as occupational or some consumer products
are not ambient air-related and as such are not included in this figure.
Figure ES-1 Conceptual model of air-related Pb exposure through inhalation
and ingestion.
The majority of Pb in the body is stored in bone (roughly 90% in adults, 70% in children; IS.5).
Much of the remaining Pb is found in soft tissues; only about 1% of Pb is found in the blood. Pb in blood
is primarily (-99%) bound to red blood cells. The small fraction of Pb in blood plasma (<1% of Pb in
blood) may be the more toxicologically active fraction of the circulating Pb. The relationship between Pb
in blood and plasma is approximately linear at relatively low daily Pb intakes and at blood Pb
concentrations below -20-30 (ig/dL. Both Pb uptake to and elimination from soft tissues are much faster
than they are in bone. Pb accumulates in bone regions undergoing the most active calcification at the time
of exposure. Pb in bone becomes distributed in trabecular (e.g., patella) and the denser cortical bones
(e.g., tibia).
Blood Pb is the most common biomarker of Pb exposure in epidemiologic studies of Pb health
effects. Overall, blood Pb levels (BLLs) have been decreasing among U.S. children and adults for the past
45 years. For children aged 1-5 years, the 1976-1980 National Health and Nutrition Examination Survey
(NHANES) showed a geometric mean BLL of 15.2 (95% CI: 14.3, 16.1) (ig/dL with nearly all children
(99.8%) exceeding 5 (ig/dL. By 2011-2016, geometric mean levels declined to 0.8 (95% CI: 0.8, 0.9)
(ig/dL with only 1.3% exceeding 5 (ig/dL (IS.6). Other common Pb exposure metrics used in
epidemiologic studies are Pb in bone, which generally reflects cumulative exposure over long periods
(months to years), and Pb in cord blood, which is an indicator of prenatal and neonatal blood Pb
concentration.
ES-6
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Blood Pb is dependent on the recent exposure history of the individual, as well as the long-term
exposure history that determines total body burden and the amount of Pb stored in the bone. The
contribution of bone Pb to blood Pb changes throughout an individual's lifetime and depends on the
duration and intensity of the exposure, age, and various other physiological stressors (e.g., nutritional
status, pregnancy, menopause, extended bed rest, hyperparathyroidism) that may affect bone remodeling,
which continuously occurs under normal conditions. In children, blood Pb is both an index of recent
exposure and, potentially, body burden, largely due to faster exchange of Pb to and from bone of children
relative to adults. Generally, bone Pb is an index of cumulative exposure and body burden. As described
previously, Pb is sequestered in two types of bone compartments: Pb in cortical bone, which is denser and
has a slower turnover rate, is a better marker of cumulative exposure than Pb in the more highly perfused
trabecular bone, which is more likely to be correlated with blood Pb concentration. During pregnancy, Pb
is transferred from the mother to the fetus. Transplacental transfer of Pb may be facilitated by an increase
in the plasma/blood Pb concentration ratio during pregnancy. Maternal-to-fetal transfer of Pb appears to
be related partly to the mobilization of Pb from the maternal skeleton.
ES.6 Health and Welfare Effects of Pb Exposure
The subsequent sections summarize the current evidence and causality determinations for health
and welfare effects in this ISA. These causality determinations appear in Figure ES-2 and Figure ES-3,
and are more fully discussed in the Integrated Synthesis and the respective health (Appendices 3-10) and
welfare effects (Appendix 11) appendices: https://asscssmcnts.cpa.go\/isa/documcnt/&dcid=359536.
ES.6.1 Health Effects of Pb Exposure
Pb exposure can disrupt important physiological pathways, triggering responses such as increased
oxidative stress and inflammation, and lead to a diverse array of health effects. In this ISA, the body of
evidence from toxicological and epidemiologic studies is evaluated for health effects that vary in severity
from minor subclinical effects to more serious effects that can lead to death. The integration of evidence
from these health studies, supported by the evidence from atmospheric chemistry, exposure assessment,
toxicokinetics, and exposure biomarker studies, contributes to the causality determinations made for the
various health outcomes. Building off the conclusions from the 2013 Pb ISA, a total of thirty causality
determinations were made for health outcomes in this ISA. These determinations are summarized in
Figure ES-2.
ES-7
-------
Causality Determinations for Health Effects of Pb
Health Outcomes
2024 Pb ISA
Nervous System Effects Ascertained During Childhood, Adolescent, and Young Adult Lifestages
Cognitive Effects
Externalizing Behaviors: Attention, Impulsivity, and Hyperactivity
Externalizing Behaviors: Conduct Disorders, Aggression, and Criminal Behavior
Internalizing Behaviors: Anxiety and Depression
Motor Function
Sensory Function
1
Social Cognition and Behavior
+
Nervous System Effects Ascertained During Adult Lifestages
Cognitive Effects
t
Psychopathological Effects
Sensory Function
Neurodegenerative Disease
t
Cardiovascular Effects
Cardiovascular Effects and Cardiovascular-Related Mortality |
Renal Effects
Renal Effects
t
Immune System Effects
Immunosuppression
Sensitization and Allergic Response
\
Autoimmunity and Autoimmune Disease
Hematological Effects
Hematological Effects |
Reproductive and Developmental Effects
Pregnancy and Birth Outcomes
Development
Female Reproductive Function
Male Reproductive Function
Effects on Other Organ Systems
Hepatic Effects
t
Metabolic Effects
+
Gastrointestinal Effects
Endocrine System Effects
Musculoskeletal Effects
Effects on Ocular Health
Respiratory Effects
Total (Nonaccidental) Mortality
Total (Nonaccidental) Mortality |
Cancer
| Cancer
1
| Causal (9) ~ Likely Causal (9) ~ Suggestive (6) ~ Inadequate (6)
+ New Causality Determination (3) Tori Change in Causality Determination since 2013 ISA (8)
Notes: (1) The 2013 Pb ISA made four causality determinations with respect to cardiovascular disease (CVD), including BP and
hypertension (causal), subclinical atherosclerosis (suggestive), coronary heart disease (CHD; causal), and cerebrovascular disease
(inadequate). This ISA follows the precedent set by the 2019 Particulate Matter and 2020 Ozone ISAs (U.S. EPA, 2020. 2019) by
making a single causality determination for cardiovascular effects. (2) The 2013 Pb ISA evaluated studies of all-cause mortality
together with studies examining cardiovascular mortality and did not issue a separate causality determination for total mortality.
Figure ES-2 Summary of causality determinations by health outcome.
ES-8
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Recent evidence continues to support causal relationships between exposure to Pb and cognitive
function decrements in children, externalizing behaviors (i.e., attention, impulsivity, and hyperactivity) in
children, cardiovascular effects and cardiovascular-related mortality, effects on development, and effects
on male reproductive function. Expanded evidence also supports causal relationships between Pb
exposure and renal effects, cognitive function decrements in adults, and total (nonaccidental) mortality.
The evidence summarized in the 2013 Pb ISA was "suggestive of a causal relationship" between Pb
exposure and renal effects and supported a "likely to be causal relationship" between Pb exposure and
cognitive effects in adults. There was no causality determination for the relationship between Pb exposure
and mortality in the 2013 Pb ISA. Evidence supporting causal relationships does not indicate a threshold
for the observed effects across the range of BLLs examined (outcome specific mean BLL ranges are
provided in Section IS.7.3 and throughout the health effects appendices). Recent evidence also indicates
that Pb exposure is likely to cause conduct disorders, internalizing behaviors, and motor function
decrements in children; depression and anxiety in adults; as well as effects on female reproductive
function, effects on pregnancy and birth outcomes, immunosuppression, musculoskeletal effects, and
cancer. Additional evidence is suggestive of a causal relationship between Pb exposure and sensory
function decrements and effects on social cognition and behavior in children; sensory function
decrements in adults; neurodegenerative disease; sensitization and allergic response; and hepatic effects;
though there are more uncertainties associated with the interpretation of the evidence for these effects.
ES.6.1.1 Effects of Pb Exposure on Health Outcomes Ascertained in Children,
Adolescents, and Young Adults
While Pb affects nearly every organ system, the nervous system appears to be one of the most
sensitive targets. Epidemiologic studies conducted in diverse populations continue to demonstrate the
harmful effects of Pb exposure on neurodevelopment in children. Given their limited exposure histories,
neurodevelopmental effects observed in young children are among the effects best substantiated as
occurring at the lowest BLLs. Specifically, blood Pb-associated effects on cognitive function are
supported by studies in populations of children (ages 4-10) with mean or group BLLs - measured
concurrently or earlier - in the range of 2-8 (ig/dL (ES.7.1.3). Notably, evidence suggests that some Pb-
related cognitive effects and neurodevelopmental effects may persist into adulthood (U.S. EPA. 2013). In
addition to cognitive effects, epidemiologic studies also demonstrate that Pb exposure is associated with
decreased attention, and increased impulsivity and hyperactivity in children (i.e., externalizing behaviors).
A small number of recent studies also serve to extend the lower bound of the mean BLLs that were
observed to be associated with inattention, impulsivity, and hyperactivity in the 2013 Pb ISA. These
prospective studies with mean maternal and cord blood Pb levels <5 (ig/dL report associations with some
measures of inattention and impulsivity. The neurodevelopmental epidemiologic evidence is supported by
findings in animal studies demonstrating both analogous effects and biological plausibility at relevant
exposure levels.
ES-9
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Pb exposure can also exert harmful effects on blood cells and blood producing organs (potentially
leading to anemia in children) and is likely to cause an increased risk of symptoms of depression and
anxiety and withdrawn behavior (i.e., internalizing behaviors), decreases in motor function, delayed
pubertal onset, as well as conduct disorders in children and young adults. There is continued uncertainty
about the timing, frequency, and duration of Pb exposures contributing to the BLLs and effects observed
in epidemiologic studies, though these uncertainties are greater in studies of older children and adults than
in studies of young children (ES.7.1.4). Despite these uncertainties, there is clear and consistent evidence
that Pb exposure leads to negative health effects in children; further, recently available evidence does not
provide evidence of a threshold for the observed neurodevelopmental effects across the range of BLLs
examined (ES.7.1.3).
ES.6.1.2 Effects of Pb Exposure on Health Outcomes Ascertained in Adults
Recent experimental animal and epidemiologic studies expand an already large body of evidence
that demonstrates the effect of Pb exposure on the cardiovascular and renal systems. The evidence most
strongly contributing to a causal relationship between Pb exposure and cardiovascular effects includes
studies reporting Pb-related increases in blood pressure, hypertension, and cardiovascular mortality. The
extent to which the effects of Pb on the cardiovascular system are reversible is not well-characterized.
Recent evidence also addresses uncertainties related to reverse causality in studies examining the renal
effects of Pb exposure and provides strong support for Pb-induced kidney dysfunction that is independent
of baseline renal function. The cardiovascular and renal effects evidence, which includes coherence of
results from epidemiologic and animal toxicological studies, is also supported by animal toxicological
evidence providing biological plausibility for the observed health effects. In particular, Pb effects on the
renin-angiotensin system provide a biologically plausible pathway through which Pb is capable of
eliciting health effects in both organ systems.
Consistent with the evidence demonstrating blood and bone Pb-associated cardiovascular
mortality, recent studies also report that Pb exposure is associated with total (nonaccidental) mortality.
The strongest supporting evidence for Pb effects on mortality comes from studies of cardiovascular
effects, which provide extensive epidemiologic and experimental animal evidence indicating pathways by
which exposure to Pb could plausibly progress from initial events to events that could lead to
cardiovascular mortality, including exacerbation of ischemic heart disease and potential myocardial
infarction. There is also very limited evidence that Pb exposure may contribute to other causes of
mortality, including Alzheimer's disease and infection, although this evidence is less established and has
greater uncertainties.
Pb exposure can also lead to cognitive function decrements, symptoms of depression and anxiety,
and immune effects in adults. Notably, the frequency, timing, level, and duration of Pb exposure causing
the effects observed in adults remains an uncertainty in the evidence, and higher past exposures may
ES-10
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contribute to the development of health effects measured later in life. Despite these uncertainties, there is
clear and consistent evidence that Pb exposure can result in harm to an array of organ systems that is
evident in adulthood, with the strongest evidence for effects on the cardiovascular and renal systems, as
well as cognitive function in adults.
ES.6.2 Welfare Effects of Pb Exposure
Several effects are associated with Pb exposure in terrestrial and aquatic organisms. Although Pb
is present in the natural environment, it has no known biological function in plants or animals. The
atmosphere and terrestrial and aquatic ecosystems are interconnected, with transfer of Pb taking place
between each of these systems (IS.2). Uptake of Pb from soils, water, sediment, and biota (via diet),
subsequent bioaccumulation, and toxicity vary greatly between biological species and across taxa, as
characterized in the 1977 AQCD (U.S. EPA. 1977). the 1986 Pb AQCD (U.S. EPA. 1986a). the 2006 Pb
AQCD (U.S. EPA. 2006). the 2013 Pb ISA (U.S. EPA. 2013). and further supported in this ISA. As
reported in the 2013 Pb ISA and preceding Pb AQCDs, effects of Pb are observed across endpoints
common to terrestrial, freshwater, and saltwater organisms. Those endpoints include reproduction,
growth, survival, neurobehavioral and hematological effects, and physiological stress, and occur at
multiple scales of biological organization, from the cellular to the ecosystem. For ecological endpoints in
this ISA, biochemical (e.g., enzymes and stress markers) responses at the suborganism level of biological
organization are grouped under the broad endpoint of "physiological stress." The effects of Pb at the
subcellular and cellular level may lead to effects on organism reproduction, growth, and survival. Effects
on these endpoints, in turn, have the potential to alter population, community, and ecosystem levels of
biological organization.
In the 2013 Pb ISA, a series of causality determinations were made for effects of Pb on plants,
invertebrates, and vertebrates in terrestrial, freshwater, and saltwater ecosystems using biological scale as
an organizing principle (U.S. EPA. 2013). Evidence published since that time supports or slightly
expands the evidence for causality in endpoints that were already established as causal in the 2013 Pb ISA
(Figure ES-3). A few studies report effects at lower concentration of Pb than in the 2013 Pb ISA. New
evidence for terrestrial (IS.8.1) and freshwater (IS.8.2) biota continues to support the existing causality
determinations from the 2013 Pb ISA and there are no changes to those causality determinations. At the
time of the 2013 Pb ISA, there were fewer studies on effects of Pb in saltwater biota than on terrestrial
and freshwater organisms, and evidence was inadequate to infer causality relationships for many
endpoints. Specifically, chronic toxicity data were lacking, and relatively few laboratory studies measured
Pb concentration in the exposure water or sediment. Since the 2013 Pb ISA, several newly available
studies verify Pb concentrations analytically and report effects on endpoints at lower concentrations than
previously observed for saltwater biota; some of these studies are chronic exposure bioassays (IS.8.3).
This additional information supports a change in causality determinations for three endpoints for saltwater
organisms (IS.8.4). Specifically, the evidence is sufficient to conclude there is a likely to be causal
ES-11
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relationship between Pb exposure and reproductive and developmental effects in saltwater invertebrates.
Additionally, the evidence is suggestive of, but not sufficient to infer, a causal relationship between Pb
exposure and saltwater vertebrate survival, and, the evidence is suggestive of, but not sufficient to infer, a
causal relationship between Pb exposure and saltwater community and ecosystem effects (Figure ES-3).
Causality Determinations for Ecological Effects of Pb
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ES.6.2.1 Effects on Development and Reproduction
Evidence from invertebrate and vertebrate studies in the Pb AQCDs, the 2013 Pb ISA, and this
ISA indicates that Pb affects reproductive performance in multiple species (IS.8.4.6). Various endpoints
measured in multiple taxa of terrestrial and aquatic organisms show impaired reproduction or
development following Pb exposure. Decreased reproduction at the organism level of biological
organization can result in a decline in how widespread a species is, the disappearance of populations of a
species, a decline in the variety of different species present, and changes in the mixture of species seen in
an ecological community. For freshwater invertebrates, recent evidence further supports previous
observations of Pb effects on reproductive endpoints at low concentrations in some sensitive species of
snails as well as zooplankton, such as cladocerans (group of small aquatic invertebrates belonging to the
subphylum Crustacea) and rotifers (small aquatic invertebrates that constitute the phylum Rotifera),
especially under chronic exposure scenarios (IS.8.4.6 and see Appendix 11. Table 11-5). Since the 2013
Pb ISA, the evidence base for Pb effects on reproductive and developmental endpoints in saltwater
invertebrates has expanded, primarily due to multiple new embryo-larval developmental assays in
mollusks and sea urchins (IS.8.4.6 and see Appendix 11. Table 11-7). This new evidence augments the
previous causality determination from the 2013 Pb ISA of suggestive of, but not sufficient to infer, a
causal relationship. This ISA concludes there is a likely to be causal relationship between Pb exposure
and reproductive and developmental effects in saltwater invertebrates.
ES.6.2.2 Effects on Growth
As reported in this ISA, the 2013 Pb ISA, and the Pb AQCDs, exposure to Pb has been shown to
have detrimental effects on growth in plants and in some species of invertebrates and vertebrates
(IS.8.4.5). Evidence for effects of Pb on growth is strongest in terrestrial plants. Evidence accumulated
over several decades of research shows that Pb inhibits photosynthesis and respiration in terrestrial plants,
both of which in turn reduce growth (U.S. EPA. 2013. 2006. 1977). Effects reported in plants largely
occur at concentrations that greatly exceed Pb concentrations typically found in U.S. soils and surface
waters, but with studies that include multiple concentrations of Pb showing increased response with
increasing Pb in water, sediment, or soil. Evidence for detrimental effects of Pb on growth in
invertebrates has been gathered most extensively in freshwater species, with growth inhibition in a few
sensitive species occurring in the range of Pb concentration values available for U.S. surface waters. In
general, juvenile organisms are more sensitive than adults. Data on growth effects of Pb in vertebrates is
limited. Causality determinations for growth in terrestrial, freshwater, and saltwater organisms remain
unchanged from the 2013 Pb ISA (Figure ES-3).
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ES.6.2.3 Effects on Survival
Survival (IS.8.4.4) may have a direct impact on population size and can lead to effects at the
community and ecosystem levels of biological organization. Pb has generally not been found to affect
survival of aquatic or terrestrial plants at concentrations found in the environment away from stationary
sources. Freshwater invertebrates are generally more sensitive to Pb exposure than other types of
organisms, with survival reduced in laboratory studies of a few species at concentrations occasionally
encountered in the environment. Studies of some freshwater invertebrates reported in the 2006 Pb AQCD
and 2013 Pb ISA indicate decreased survival at <20 |ig Pb/L under some water quality conditions. Several
studies since the 2013 Pb ISA provide further characterization for known effects on survival in a few
sensitive species of freshwater invertebrates, notably snails and amphipods (shrimp-like crustaceans), at
analytically verified chronic exposure <15 |ig Pb/L (IS.8.4.4). Limited studies with vertebrates showed
adverse effects of Pb on survival at concentrations higher than typical Pb levels in the environment,
although juvenile organisms are usually more sensitive than adults. The 2013 Pb ISA causality
determination for survival in saltwater vertebrates was inadequate. Additional evidence in this ISA
(IS.8.4.4) from laboratory-based bioassays in a few saltwater fish species in which Pb exposure
concentration was analytically verified demonstrates effects on survival in chronic exposures to Pb
(Appendix 11. Table 11-7). Based on these new chronic studies in saltwater fish, the evidence is
suggestive of, but not sufficient to infer, a causal relationship between Pb exposure and saltwater
vertebrate survival.
ES.6.2.4 Neurobehavioral Effects
Pb is a known to cause impairments in the nervous system of invertebrates and vertebrates.
Historical and recent evidence of Pb effects on terrestrial and freshwater animals indicates that Pb
adversely affects behaviors, such as food consumption, locomotion, behavioral regulation of body
temperature, and prey capture. Additional evidence since the 2013 Pb ISA includes studies quantifying
alterations in foraging and feeding behavior in bees and changes in locomotion in freshwater amphipods,
bivalves, and zebrafish (IS.8.4.3). The causality determinations for neurobehavioral effects of Pb in
terrestrial, freshwater, and saltwater organisms remain unchanged from the 2013 Pb ISA (Figure ES-3).
ES.6.2.5 Hematological Effects
As reported in the Pb AQCDs and 2013 Pb ISA, hematological effects of Pb exposure in wildlife
include inhibition of S-aminolevulinic acid dehydratase (ALAD; an important rate-limiting enzyme
needed for heme production) and altered blood cell counts and serum profiles. Decreased ALAD activity
is commonly recognized as an indicator of Pb exposure across a wide range of animals as shown in both
field and laboratory studies. Previous studies have indicated considerable species differences in ALAD
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activity in response to Pb. Since the 2013 Pb ISA, new studies in terrestrial birds, amphibians, and
mammals have continued to support the connection between Pb exposure and hematological effects
(IS.8.4.2). In contrast, fewer studies were identified that quantified ALAD response in terrestrial or
freshwater invertebrates, freshwater vertebrates, or in saltwater organisms. The causality determinations
for hematological effects of Pb in biota remain unchanged from the 2013 Pb ISA (Figure ES-3).
ES.6.2.6 Effects on Physiological Stress
Increased levels of antioxidant enzymes (in response to oxidative stress or altered cell signaling)
and increased lipid peroxidation (the process by which free radicals induce the oxidation of fatty acids,
leading to cell membrane damage) are reliable biomarkers of various stresses. Oxidative damage and
antioxidant activity have been observed in field studies in a wide range of species in terrestrial and
aquatic environments when Pb is present (often along with other chemical stressors), and also following
laboratory exposures to Pb without other stressors in plants, invertebrates and vertebrates (IS.8.4.1).
Changes in markers of physiological stress may indicate increased susceptibility to other stressors, as well
as diminished fitness of individual organisms. Causality determinations for physiological stress in
terrestrial, freshwater, and saltwater organisms remain unchanged from the 2013 Pb ISA (Figure ES-3).
ES.6.2.7 Community and Ecosystem Effects
Uptake of Pb by terrestrial and aquatic organisms and subsequent adverse effects on survival,
growth, development, and reproduction at the organism level can sometimes lead to effects at higher
levels of biological organization including populations, communities, and ecosystems. In terrestrial
habitats, soil microbial, plant, and animal communities may be affected in locations where soil Pb
concentration is elevated, such as in the proximity to historic metal extracting and processing point
sources. In freshwater ecosystems, shifts in sediment-associated microbial and invertebrate communities
and aquatic plant communities are linked to the presence of Pb as well as other stressors. For terrestrial
and freshwater systems, the likely to be causal determinations remain unchanged from the 2013 Pb ISA.
For saltwater ecosystems, new experimental and observational studies have examined the relationship
between Pb in sediment, and microbial abundance and/or diversity and saltwater foraminifera (single-
celled marine organisms, usually with shells) communities (IS.8.4.7). These studies show that diversity
and distribution of these organisms varies with Pb concentration and co-stressors in the environment and
at different locations. This new evidence is suggestive of a causal relationship between Pb exposure and
saltwater community and ecosystem effects which is a change from the 2013 Pb ISA. Although the
presence of Pb is associated with shifts in biological communities, this metal rarely occurs as a sole
contaminant in natural systems, making the contribution of Pb to the observed effects difficult to isolate
in many locations. Furthermore, the variability of conditions in the environment affects Pb bioavailability
and organism response making it difficult to characterize effects of Pb at the ecosystem scale.
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ES.7 Key Aspects of Health and Welfare Effects Evidence
In addition to causality determinations, this ISA also reaches conclusions on other policy-relevant
topics. These conclusions are drawn from a careful evaluation of the available evidence and the extent to
which recent studies have addressed or reduced uncertainties from previous assessments. Conclusions on
key policy-relevant topics are summarized below.
ES.7.1 Health Effects Evidence: Key Findings
In addition to the causality determinations for health effects (ES.6.1), the evidence evaluated in
this ISA addresses some of the key policy-relevant issues of this NAAQS review, as outlined in Volume 2
of the Pb IRP (U.S. EPA. 2022a). A summary of this health evidence and Pb ISA conclusions is provided
below and discussed in more detail in the Integrated Synthesis and supporting appendices.
ES.7.1.1 At-Risk Populations
The NAAQS are intended to protect public health with an adequate margin of safety, including
protection for those potentially at increased risk for health effects in response to exposure to a criteria air
pollutant [e.g., Pb; see Preamble (U.S. EPA. 2015)1. In addition to consideration of Pb-related health
effects observed among populations with diverse characteristics, this ISA also considers those studies that
examine specific populations or lifestages that may be at increased risk of Pb-related health effects, using
a pragmatic approach to characterize the strength of the evidence (U.S. EPA. 2015). The risk of health
effects from exposure to Pb may be modified by intrinsic (e.g., pre-existing disease, genetic factors) or
extrinsic (e.g., sociodemographic or behavioral factors) factors, differences in internal dose, or differences
in exposure to Pb in the environment. While a combination of factors (e.g., residential location and SES)
may increase the risk of Pb-related health effects in portions of the population, information on the
interaction among factors remains limited. Thus, this ISA characterizes the individual factors that
potentially result in increased risk for Pb-related health effects [see Preamble (U.S. EPA. 2015)1. There is
adequate evidence to classify children, minority populations, individuals in close proximity to Pb sources,
individuals living in residences with factors contributing to increased house dust Pb levels, individuals
with certain genetic variants, individuals with high stress levels, and those with certain nutritional
excesses or deficiencies as populations at increased risk to the health effects of Pb exposure (IS.7.4).
These conclusions are based on the consistency in findings across studies, as well as on coherence of
results from different scientific disciplines. There is suggestive evidence for several other factors
contributing to potentially increased risk of Pb-related health effects: older age, sex, pre-existing diabetes,
low socioeconomic status, and high levels of exposure to other metals.
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ES.7.1.2 Air-Pb-to-Blood-Pb Relationships
The relationship between air Pb and blood Pb is commonly characterized as a "slope factor,"
which describes the incremental change in blood Pb levels relative to a change in air Pb concentrations
(IS.9.1). A larger slope indicates a larger estimated incremental contribution of air Pb to the blood Pb
level in exposed populations. Epidemiologic studies evaluating air-to-blood slope factors include various
study locations, populations, and analytic methodologies (e.g., model form and other considerations, such
as soil Pb, that are accounted for in the model), all of which contribute to variation in the estimated
slopes. Results described in the 2013 Pb ISA (U.S. EPA. 2013) provide a range of air-to-blood slope
estimates from 4 to 9 (ig/dL per (ig/m3 in studies of children. Newer studies after the phaseout of leaded
gasoline and not focused on communities near significant air Pb sources show increasing slope factors
with decreasing air Pb concentrations.
ES.7.1.3 Concentration-Response Relationships for Human Health Effects
In assessing the relationship between Pb exposure and human health effects, evidence from each
previous assessment (U.S. EPA. 2013. 2006) demonstrates that progressively lower BLLs or Pb
exposures are associated with cognitive deficits in children. The evidence assessed in the 2013 Pb ISA
found that cognitive effects in children were substantiated to occur in populations with mean BLLs
between 2 and 8 (ig/dL. Recent studies generally include somewhat older children or employ modelling
strategies designed to answer relatively narrow research questions and consequently do not have the
attributes of the studies on which the conclusion of the 2013 Pb ISA was based (i.e., early childhood
BLLs, consideration of peak BLLs, or concurrent BLLs in young children). Therefore, the recently
available studies were not designed and may not have the sensitivity to detect the effect or hazard at these
very low BLLs, nor do they provide evidence of a threshold for the effect across the range of BLLs
examined.
Compelling evidence in the 2013 Pb ISA also supported a larger incremental negative effect of
Pb on children's IQ at lower BLLs compared to higher BLLs (for BLLs ranging from 2.5 to 33.2 (ig/dL;
Section IS.9.2). Only a few recent studies evaluate the shape of the concentration-response (C-R) function
for the relationship between Pb exposure and cognitive effects in children, but recent evidence continues
to support the conclusions from the 2013 Pb ISA. Possible explanations specific to nonlinear relationships
observed in studies of Pb exposure in children include a smaller incremental effect at higher Pb
concentrations due to covarying risk factors, though the evidence does not reveal a consistent set of
covarying risk factors that explain the differences in the blood Pb-IQ C-R relationship observed in
epidemiologic studies. Additionally, although evidence indicates a larger incremental effect of Pb
exposure on IQ at lower BLLs, consistent findings of higher mean IQ at lower BLLs indicate that the
absolute magnitude of the effect of Pb exposure on cognitive function declines with decreasing BLL.
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ES.7.1.4 Lifestages and Timing of Pb Exposure Contributing to Observed Nervous
System Effects
As discussed in Section ES.5, blood Pb may reflect recent as well as past exposures because Pb is
both taken up by and released from the bone. The resulting uncertainty regarding the relative proportion
of blood Pb from recent versus past exposure is greater in adults and older children than in young children
who have shorter exposure histories. As a result, there is inherent uncertainty in the level, timing,
frequency, and duration of Pb exposures contributing to associations between adult and older children's
BLLs and health outcomes observed in epidemiologic studies. In epidemiologic studies of nervous system
effects with BLLs measured in younger children, recent evidence is consistent with findings from the
2013 Pb ISA, which consistently showed that BLLs measured during various lifestages and time periods
(i.e., prenatal, early childhood, childhood average, and concurrent with the outcome) were associated with
nervous system effects in children. A notable uncertainty in the interpretation of this evidence is the
typically high correlation between blood Pb measurements at different ages in childhood, making it
difficult to discern the relative importance of the various exposure metrics (i.e., BLLs at different ages)
used in epidemiologic studies. Nonetheless, the epidemiologic evidence is supported by experimental
evidence in monkeys that indicates that Pb exposures during multiple lifestages and time periods,
including prenatal only, prenatal plus lactational, postnatal only, or lifetime starting during the juvenile
period, induce impairments in cognitive function when assessed between ages 6 and 10 years.
Additionally, recent prospective epidemiologic studies observed associations between childhood BLLs
and decrements in IQ during late adolescence (18-19 years) and mid-adulthood (38-45 years). These
findings provide insight into the persistence of Pb-associated cognitive function decrements and are
consistent with the understanding that the nervous system continues to develop throughout childhood and
into adolescence.
ES.7.2 Welfare Effects Evidence: Key Findings
Effects of Pb in ecosystems are primarily associated with Pb from deposition and other sources,
subsequent transport, and exposure through environmental media (soil, water, sediment, biota). Pb
bioaccumulates in plants and animals in terrestrial, freshwater, and saltwater environments; however, the
relative contribution of Pb from different sources is usually not known. The share of Pb effects attributed
to atmospheric sources is difficult to quantitatively assess due to limited information on Pb deposition to
soils, water, and sediments, a lack of Pb-apportionment studies in biota, and kinetics of Pb distribution
within organisms in long-term exposure scenarios.
Exposure of organisms to Pb can be via one or more pathways (e.g., uptake from soil or water,
ingestion). For Pb to interact with a biological membrane and be taken up into an organism it must be
bioavailable (IS.8). Generally, the greater amount of Pb available as the free Pb ion, the greater
bioavailability. Conditions in the environment, such as soil composition and soil and water chemistry,
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modify Pb bioavailability and subsequent toxicity to organisms. Once Pb uptake occurs, a variety of
effects may occur in organisms, including impaired reproduction, decreased growth, and reduced survival,
as documented in this ISA, the 2013 Pb ISA and the Pb AQCDs. These effects on individual organisms
may lead to effects at the population, community, and ecosystem level of biological organization. In both
terrestrial and aquatic organisms, gradients in response are observed with increasing concentration of Pb
in laboratory and field studies. However, the level at which Pb elicits a specific effect in a natural system
is difficult to establish due to the influence of other environmental variables (e.g., pH and organic matter)
on both Pb bioavailability and toxicity, and also because of substantial species differences in Pb
sensitivity. Some laboratory studies report effects within the range of Pb detected in environmental media
over the past several decades. Specifically, effects on reproduction, growth, and survival in sensitive
freshwater invertebrates are well-characterized from controlled studies at concentrations at or near Pb
concentrations occasionally encountered in U.S. fresh surface waters. There are considerable uncertainties
associated with generalizing effects observed in controlled studies on a single species to effects at higher
levels of biological organization. Furthermore, available studies on community and ecosystem-level
effects are usually from heavily contaminated areas where Pb concentrations are much higher than
typically encountered in the general environment. Measurements of the contribution of atmospheric Pb to
specific sites that are not directly contaminated by a known point source are generally unavailable, and
the connection between air concentration of Pb and ecosystem exposure continues to be poorly
characterized, as was reported in the 2013 Pb ISA.
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ES.8 References
U.S. EPA (U.S. Environmental Protection Agency). (1977). Air quality criteria for lead [EPA Report]. (EPA-600/8-
77-017). Washington, DC. http://nepis.epa.gov/Exe/ZvPURL.cgi?Dockev=20013GWR.txt.
U.S. EPA (U.S. Environmental Protection Agency). (1986a). Air quality criteria for lead: Volume I of IV [EPA
Report]. (EPA-600/8-83/028aF). Research Triangle Park, NC.
http://cfpub.epa.gov/ncea/cfm/recordisplav.cfm?deid=32647.
U.S. EPA (U.S. Environmental Protection Agency). (1986b). Lead effects on cardiovascular function, early
development, and stature: An addendum to U.S. EPA Air Quality Criteria for Lead (1986) [EPA Report].
(EPA-600/8-83/028aF). Washington. DC.
U.S. EPA (U.S. Environmental Protection Agency). (1990). Air quality criteria for lead: Supplement to the 1986
addendum [EPA Report]. (EPA/600/8-89/049F). Washington, DC.
https://nepis.epa. gov/Exe/ZvPDF.cgi?Dockev=30001IKR.PDF.
U.S. EPA (U.S. Environmental Protection Agency). (2006). Air quality criteria for lead [EPA Report]. (EPA/600/R-
05/144aF-bF). Research Triangle Park, NC. https://cfpub.epa.gov/ncea/risk/recordisplav.cfm?deid=158823.
U.S. EPA (U.S. Environmental Protection Agency). (2013). Integrated science assessment for lead [EPA Report].
(EPA/600/R-10/075F). Washington, DC. https://nepis.epa.gov/Exe/ZvPURL.cgi?Dockev=P100K82L.txt.
U.S. EPA (U.S. Environmental Protection Agency). (2015). Preamble to the Integrated Science Assessments [EPA
Report]. (EPA/600/R-15/067). Research Triangle Park, NC: U.S. Enviromnental Protection Agency, Office
of Research and Development, National Center for Enviromnental Assessment, RTP Division.
https ://cfpub .epa. gov/ncea/isa/recordisplay. cfm?deid=310244.
U.S. EPA (U.S. Environmental Protection Agency). (2019). Integrated Science Assessment (ISA) for particulate
matter (final report, Dec 2019). (EPA/600/R-19/188). Washington, DC.
https ://cfpub .epa. gov/ncea/isa/recordisplav. cfm?deid=347534.
U.S. EPA (U.S. Environmental Protection Agency). (2020). Integrated Science Assessment (ISA) for ozone and
related photochemical oxidants (final report, Apr 2020) [EPA Report]. (EPA/600/R-20/012). Washington,
DC. https://nepis.epa.gov/Exe/ZvPURL.cgi?Dockev=P101 llKI.txt.
U.S. EPA (U.S. Environmental Protection Agency). (2021). 2017 National Emissionslinventory: January 2021
updated release, technical support document [EPA Report]. (EPA-454/R-21-001). Research Triangle Park,
NC. https://www.epa.gov/sites/default/files/2021-02/documents/nei2017 tsd full ian2021.pdf.
U.S. EPA (U.S. Environmental Protection Agency). (2022a). Integrated review plan for the National Ambient Air
Quality Standards for lead. Volume 2: Planning for the review and the Integrated Science Assessment
[EPA Report]. (EPA-452/R-22-003b). Research Triangle Park, NC: U.S. Enviromnental Protection
Agency, Office of Air Quality Planning and Standards and Office of Research and Development.
https://nepis.epa.gov/Exe/ZvPURL.cgi?Dockev=P10148PX.txt.
U.S. EPA (U.S. Environmental Protection Agency). (2022b). Overview of lead (Pb) air quality in the United States,
2021. Washington, DC.
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INTEGRATED SYNTHESIS FOR LEAD
Overall Conclusions of the Lead (Pb) Integrated Science Assessment (ISA)
Human Health Effects
• Recent studies support and expand upon the strong body of evidence spanning scientific disciplines, reaffirming
causal relationships between Pb exposure and several nervous system effects in children, including cognitive
function decrements and externalizing behaviors (i.e., attention, impulsivity, and hyperactivity).
• Recent evidence also continues to support causal relationships between Pb exposure and a number of other
health effects, including cardiovascular effects and cardiovascular-related mortality, hematological effects,
developmental effects, and effects on male reproductive function.
• Expanded evidence supports causal relationships between Pb exposure and (1) renal effects (previously
suggestive of a causal relationship) and (2) cognitive function in adults (previously likely to be causal).
• Drawing largely on evidence for Pb-related cardiovascular mortality, a new causality determination affirms a
causal relationship between Pb exposure and total (nonaccidental) mortality.
• Recent experimental and epidemiologic evidence supports likely to be causal relationships between Pb
exposure and conduct disorders in children and young adults, internalizing behaviors in children and
adolescents, motor function decrements in children, psychopathological effects in adults, immunosuppression,
musculoskeletal effects, effects on female reproductive function, effects on pregnancy and birth outcomes, and
cancer.
• For all other health effect categories, uncertainties and limitations in the scientific evidence contribute to
causality determinations that the evidence is suggestive of, but not sufficient to infer, a causal relationship or
inadequate to infer the presence or absence of a causal relationship.
• Many population subgroups and different lifestages have been shown to be at increased risk of Pb-related health
effects resulting from variation in exposure or biological responses to exposure. Among populations and
lifestages evaluated in this ISA, current scientific evidence is adequate to conclude that children, people living
in proximity to Pb sources, people with specific genetic variants, people with increased stress, and populations
with certain nutritional or residential factors may be at disproportionate risk for Pb-related health effects. There
is suggestive evidence that older age, sex, pre-existing disease, socioeconomic status (SES), and exposure to
other metals may increase risk for health effects of Pb exposure.
Welfare Effects
• Effects of Pb in ecosystems are primarily associated with Pb from deposition and other sources, subsequent
transport, and exposure through environmental media (soil, water, sediment, biota). Pb bioaccumulates in plants
and animals in terrestrial, freshwater, and saltwater environments; however, the relative contribution of Pb from
different sources is usually not known.
• Effects of Pb are observed in terrestrial, freshwater, and saltwater organisms across several levels of biological
organization (i.e., from the cellular level of organization through individual organisms to the level of
communities and ecosystems). Most evidence is from toxicity bioassays on individual organisms, rather than
field-based studies.
• In most cases, new research affirms the conclusions in the 2013 Pb ISA for the endpoints of physiological
stress, hematological effects, neurobehavior, survival, growth, reproduction and development, and community
and ecosystem effects in terrestrial and freshwater biota. A few studies report effects at lower concentrations
than in the 2013 Pb ISA.
• Additional studies in saltwater organisms address some of the uncertainties identified in the 2013 Pb ISA.
There is sufficient new evidence to support a likely to be causal relationship between Pb exposure and
reproductive and developmental effects in saltwater invertebrates. For two other endpoints, survival in
saltwater vertebrates (based on fish studies) and effects on saltwater communities and ecosystems, new
evidence is suggestive of, but not sufficient to infer, a causal relationship.
IS-1
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IS.1
Introduction
IS.1.1 Purpose and Overview
The Integrated Science Assessments (ISAs), prepared by the U.S. Environmental Protection
Agency (U.S. EPA), serve as the scientific foundation of the National Ambient Air Quality Standards
(NAAQS) review process.1 The ISA is a comprehensive evaluation and synthesis of the policy-relevant
science "useful in indicating the kind and extent of all identifiable effects on public health or welfare,2
which may be expected from the presence of [a] pollutant in the ambient air," as described in Section 108
of the Clean Air Act (42 U.S. Code [U.S.C.] 7408).3 For this ISA, "policy-relevant" science is described
in Volume 2 of the Integrated Review Plan (IRP) for Lead (Pb) (U.S. EPA, 2022a) as referring to
"scientific information and analyses intended to address key questions related to the adequacy of the
standards." Those "key questions" are also laid out in Volume 2 of the IRP. As stated in the Preamble to
the ISAs (U.S. EPA, 2015), hereafter "Preamble," "[t]he key policy-relevant questions included in the
IRP serve to clarify and focus the NAAQS review on the critical scientific and policy issues, including
addressing uncertainties discussed during the previous review and newly emerging literature." This ISA
reviews and synthesizes the air quality criteria for the health and welfare effects of Pb. It draws on the
existing body of evidence to evaluate and describe the current state of scientific knowledge on the most
relevant issues pertinent to the current review of the Pb NAAQS, to identify changes in the scientific
evidence since the previous review, and to describe remaining or newly identified uncertainties and
limitations in the evidence.
This Integrated Synthesis (IS) is the main body of the Pb ISA. The following sections provide a
concise synopsis of the ISA conclusions and synthesize the key findings considered in characterizing Pb
exposure and relationships with health and welfare effects. The IS includes summaries of key information
for each topic area covered in 12 appendices to the Pb ISA, including atmospheric science, sources, and
environmental distribution; exposure, biomarkers, and toxicokinetics; the nature of health and welfare
effects associated with Pb exposure, including causality determinations for relationships between
exposure to Pb and specific types of health and welfare effects; and the human lifestages and populations
at increased risk of the effects of Pb. This IS also discusses the evidence related to other policy-relevant
issues, such as the exposure durations, metrics, and concentrations eliciting health and welfare effects; the
Section 109(d)(1) of the Clean Air Act requires periodic review and, if appropriate, revision of existing air quality
criteria to reflect advances in scientific knowledge on the effects of the pollutant on public health and welfare. Under
the same provision, EPA is also to periodically review and, if appropriate, revise the NAAQS based on the revised
air quality criteria.
2Under section 302(h) of the Clean Air Act, effects on welfare include, but are not limited to, "effects on soils,
water, crops, vegetation, manmade materials, animals, wildlife, weather, visibility, and climate, damage to and
deterioration of property, and hazards to transportation, as well as effects on economic values and on personal
comfort and well-being."
3The general process for developing an ISA, including the framework for evaluating weight of evidence and drawing
scientific conclusions and causal judgments, is described in a companion document, the Preamble to the ISAs.
IS-2
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concentration-response (C-R) relationships for specific effects, including the overall shape and
discernibility of thresholds in these relationships; and the public health and welfare impact of effects
associated with exposure to Pb.
The 2024 Pb ISA will inform U.S. EPA decisions on the primary and secondary NAAQS for Pb.
The primary Pb NAAQS are established to protect public health with an adequate margin of safety,
including the health of at-risk populations such as children. The secondary Pb NAAQS are intended to
protect the public welfare from known or anticipated adverse effects associated with the presence of the
pollutant in ambient air. The current primary and secondary Pb NAAQS were established in 2008. In that
review, the levels of the primary and secondary standards were lowered tenfold, from the 1978 levels of
1.5 (ig/m3 to 0.15 |ig/nr\ The averaging time was revised from a calendar quarter average to a rolling
three-month period with a maximum (not-to-be-exceeded) form, evaluated over a three-year period. The
revised primary standard was established to protect against air Pb-related human health effects, including
intelligence quotient (IQ) loss, in the most highly exposed children. The secondary standard was set equal
to the primary standard for requisite protection of organisms and ecosystems. The most recent review of
the Pb NAAQS was completed in 2016, at which time the standards set in 2008 were retained without
revision.
IS.1.2 Pb Integrated Science Assessment Process and Development
Each NAAQS review begins with a "Call for Information" published in the Federal Register that
announces the start of the review and invites the public to assist in this process by identifying relevant
research studies in the subject areas of concern. For this review of the Pb NAAQS, the Call for
Information was published in the Federal Register on July 7, 2020 (85 FR 40641). Following the Call for
Information, the planning phase of the review includes development of an IRP, which is made available
for public comment and provided to the Clean Air Scientific Advisory Committee (CASAC) for review or
consultation. Volume 2 of the IRP for Pb addresses the general approach for the review and planning for
the ISA (U.S. EPA. 2022a).
The process for developing this ISA is described in detail in Appendix 12 of this ISA, Process for
Developing the Pb Integrated Science Assessment. Through iterative NAAQS reviews, ISAs build on
evidence and conclusions from previous assessments. The previous ISA for Pb was published in 2013
(U.S. EPA. 2013a) and included peer-reviewed literature published through September 2011. Prior Pb
assessments include the 2006 Air Quality Criteria Document (AQCD) for Pb (U.S. EPA. 2006). the 1986
Pb AQCD (U.S. EPA. 1986b) and its associated addendum (U.S. EPA. 1986d). the 1990 Supplement to
the 1986 addendum (U.S. EPA. 1990). and the 1977 AQCD for Pb (U.S. EPA. 1977). This ISA focuses
on synthesizing and integrating the evidence that has become available since the 2013 Pb ISA with the
information and conclusions from previous assessments. Important older studies from the 2013 Pb ISA or
from the Pb AQCDs may be drawn on to reinforce key concepts and conclusions. Older studies also may
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be the primary focus in some subject areas or scientific disciplines where research efforts have subsided,
and/or where these older studies remain the definitive works available in the literature. The general steps
for ISA development include literature search and study selection; evaluating study quality; developing
initial draft materials for peer-input consultation; evaluating, synthesizing, and integrating evidence; and
developing scientific conclusions and causality determinations (U.S. EPA. 2015).
These steps are described in greater detail in the Preamble (U.S. EPA. 2015). which provides a
general framework for developing ISAs, and in the Process Appendix (Appendix 12). which supplements
the Preamble with additional details specific to this ISA including methods for documentation, literature
review, study quality evaluation, public engagement, and quality assurance (QA). As described in the
Preamble, the U.S. EPA uses a structured and transparent process to evaluate scientific information and to
determine the causal nature of relationships between air pollution and health and welfare effects [see
Preamble (U.S. EPA. 2015)1. Development of the ISA includes approaches for literature searches,
application of criteria for selecting and evaluating relevant studies, and application of a framework for
evaluating the weight of evidence and forming causality determinations. As part of the external review
process, one or more drafts of the ISA are made available to the public and undergo formal review by the
CASAC, an independent scientific committee appointed by the U.S. EPA Administrator.
Studies considered in the development of the Pb ISA are documented in the U.S. EPA HERO
database. The publicly accessible HERO project page for this ISA contains the references that were
considered for inclusion and provides bibliographic information and abstracts. Within HERO, each
reference has a unique HERO ID number. References can be viewed individually or filtered by appendix,
discipline, or the draft in which they are referenced.
IS.1.2.1 Scope of the Pb ISA
Pb is a multimedia and persistent pollutant that contributes complexities to the review of the Pb
NAAQS. Pb emitted into ambient air may subsequently be found in multiple environmental media
(i.e., soil, water, sediment, biota), contributing to multiple pathways of exposure for humans and
ecological receptors. This multimedia distribution of, and multipathway exposure to, air-related Pb has a
key role in the Agency's consideration of the Pb NAAQS. The Pb ISA includes research relevant to
assessing the health and welfare effects of Pb exposure. Health effects evidence evaluated in the ISA
includes experimental animal toxicological studies and observational epidemiologic studies. Welfare
evidence included in the Pb ISA focuses specifically on ecological effects. In addition to the human
health and welfare effects of Pb, the ISA also evaluates other scientific information on sources of Pb to
ambient air, measurement, and concentrations of Pb in ambient air, fate, and transport of Pb in the
environment, pathways of human and ecological exposure, toxicokinetic characteristics of Pb in the
human body, and characterization of population exposures to Pb.
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The scope of the health portions of the ISA are explicitly defined by scoping statements that
generally characterize the parameters for study inclusion to aid in identifying the most relevant evidence.
The use of scoping statements to define study relevance is consistent with recommendations by the
National Academies of Sciences, Engineering, and Medicine for improving the design of risk assessment
through planning, scoping, and problem formulation to better meet the needs of decision makers
(NASEM. 2018). The statement used to define the scope of the health effects portion of this ISA
comprises Population, Exposure, Comparison, Outcome, and Study Design (PECOS) components. There
are discipline-specific PECOS criteria for experimental and epidemiologic studies. For experimental
studies, the scope of the evidence used for this ISA encompassed studies of nonhuman mammalian animal
species with exposures that are relevant to the range of human exposures, with mean blood Pb levels
(BLLs) up to 30 (ig/dL, which is about one order of magnitude above the 95th percentile of the 2011—
2016 National Health and Nutrition Examination Survey (NHANES) distribution of BLLs in children
(Egan et al.. 2021). The evaluation of epidemiologic studies focused on the association between exposure
to Pb (as indicated by Pb levels in blood, bone, and teeth; validated environmental indicators of Pb
exposure; or intervention groups in randomized trials and quasi-experimental studies) and an ensemble of
health effects, including effects on the nervous system, cardiovascular effects, and reproductive and
developmental outcomes. Emphasis was placed on studies conducted in non-occupationally exposed
populations, but recent longitudinal studies of occupational exposure to Pb published since the literature
cutoff date for the 2013 Pb ISA were considered insofar as they addressed a topic that was of particular
relevance to the NAAQS review (e.g., longitudinal studies designed to examine recent versus historical
Pb exposure). Additionally, the following types of health studies are generally considered to fall outside
the scope and are not included in the ISA: review articles (which typically present summaries or
interpretations of existing studies rather than bringing forward new information in the form of original
research or new analyses); Pb poisoning studies or clinical reports (e.g., involving accidental exposures to
very high amounts of Pb described in clinical reports that may be extremely unlikely to be experienced
under ambient air exposure conditions); and risk or benefit analyses (e.g., that apply existing C-R
functions or effect estimates to exposure estimates for differing cases). For the health appendices, the
PECOS statement defines the scope of the studies considered in the assessment of health evidence and
establishes study inclusion criteria thereby facilitating identification of the most relevant literature to
inform the ISA for each health discipline.
The statement used to define the scope of the ecological effects portion of this ISA comprises
Level of Biological Organization, Exposure, Comparison, Endpoint, and Study Design (LECES). The
LECES statement developed by the U.S. EPA specifically for the purpose of scoping literature for the
ISAs, was based on the PECOS with some concepts substituted to provide a better fit with ecological
science. In the LECES, "population" (PECOS) is replaced with "level of biological organization"
(LECES) and "outcome" (PECOS) is replaced with "endpoint" (LECES). The LECES statement aids in
identifying the relevant evidence in the literature for ecological effects of Pb. Other topics within scope,
in addition to Pb effects on biota described in the LECES criteria above, include effects of Pb
biogeochemistry on bioavailability in terrestrial, freshwater, and saltwater environments; subsequent
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vulnerability of particular organisms, populations, communities, or ecosystems, as well as key
uncertainties and limitations in the evidence identified in the previous review. Concentrations relevant to
the welfare effects of Pb consider the range of Pb concentrations in the environment and the available
evidence for concentrations at which effects are observed in plants, invertebrates, and vertebrates. Effects
observed at or near Pb concentrations measured in ambient soil, sediment, and water for which local
contamination is not thought to be a primary contributor are emphasized. Concentration cutoff values
were applied when evaluating the ecological literature published since the 2013 Pb ISA (Appendix 12).
For soil, the cutoff value for screening of terrestrial studies of Pb exposure and effects was set at a
concentration of approximately 230 mg Pb/kg of soil. For aqueous exposures, the cutoff value for study
screening was approximately 10 |ig Pb/L and, for sediments, the literature cutoff value for study
screening was approximately 300 mg Pb/kg dry weight (Appendix 12. Table 12-4). Studies at higher
concentrations were included only to the extent that they informed mechanisms of action, exposure-
response, or the wide range of sensitivity to Pb across taxa. Areas outside of the scope for ecological
effects in the Pb ISA included site-specific studies in non-U.S. locations that did not contribute novel
insights on Pb biogeochemistry or effects. Studies on mine tailings, biochar, industrial effluent, sewage,
ship breaking, bioremediation of highly contaminated sites, and ingestion of Pb shot, fishing tackle or
pellets were not within the scope of the ISA due to high concentration of Pb and lack of a connection to
an air-related source or process.
IS.1.2.2 Organization of the ISA
The ISA consists of the Front Matter (list of authors, contributors, and reviewers), Executive
Summary (ES), IS and 12 appendices: https://assessments.epa.gov/isa/document/&deid=359536. This IS
consolidates the key findings from the appendices considered in characterizing Pb exposure and
relationships with human and welfare effects. Subsequent appendices are organized by subject area and
include a detailed assessment and description of atmospheric science (Appendix 1), exposure
(Appendix 2), health evidence (Appendix 3-Appendix 10), welfare evidence (Appendix 11), and the ISA
development process (Appendix 12). Appendices for each broad health effect category (e.g., nervous
system effects) discuss potential biological pathways and conclude with a causality determination
describing the strength of the evidence between exposure to Pb and the outcome(s) under consideration.
Likewise, the appendix devoted to welfare evidence (Appendix 11) includes causality determinations for
multiple effects on ecosystems.
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Organization of the 2024 Pb ISA:
• Front Matter
• Executive Summary
• Integrated Synthesis
• Appendix 1. Lead Source to Concentration
• Appendix 2. Exposure, Toxicokinetics, and Biomarkers
• Appendix 3. Nervous System Effects
• Appendix 4. Cardiovascular Effects
• Appendix 5. Renal Effects
• Appendix 6. Immune System Effects
• Appendix 7. Hematological Effects
• Appendix 8. Reproductive and Developmental Effects
• Appendix 9. Effects on Other Organ Systems and Mortality
• Appendix 10. Cancer
• Appendix 11. Effects of Lead in Terrestrial and Aquatic Ecosystems
• Appendix 12. Process for Developing the Pb Integrated Science Assessment
IS.1.2.3 Quality Assurance Summary
The use of QA and peer review helps ensure that the U.S. EPA conducts high-quality science
assessments that can be used to help policymakers, industry, and the public make informed decisions. QA
activities performed by the U.S. EPA ensure that environmental data are of sufficient quality to support
the Agency's intended use. The U.S. EPA has developed a detailed Program-level QA Project Plan
(PQAPP) for the ISA Program to describe the technical approach and associated QA/quality control
procedures associated with the ISA Program. All QA objectives and measurement criteria detailed in the
PQAPP have been employed in developing this ISA. Furthermore, the Pb ISA is classified as a Highly
Influential Scientific Assessment (HISA), which is defined by the Office of Management and Budget
(OMB) as a scientific assessment that is novel, controversial, or precedent-setting, or has significant
interagency interest (Bolton. 2004). OMB requires a HISA to be peer reviewed before dissemination. To
meet this requirement, the U.S. EPA engages CASAC as an independent federal advisory committee to
conduct peer reviews. Both peer-review comments provided by the CASAC panel and public comments
submitted to the panel during its deliberations about the external review draft were considered in the
development of the final ISA. For a more detailed discussion of peer review and QA, see Appendix 12.
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IS.1.2.4 Evaluation of the Evidence
This ISA draws conclusions about the causal nature of relationships between exposure to Pb and
categories of related health and welfare effects, the concentrations at which effects are observed, and the
populations and organisms most affected by Pb, by integrating recent evidence across scientific
disciplines and building on the evidence from previous assessments. Determinations are made about
causation, not just association, and are based on judgments of nine aspects of the evidence, including
consistency, coherence, and biological plausibility of observed effects, and on related uncertainties (U.S.
EPA. 2015). In evaluating the evidence, emphasis is placed on the consideration of the strengths,
limitations, and possible roles of chance, confounding, and other biases that may affect the interpretation
and/or the strength of inference from the results of individual studies. The ISA uses a formal causal
framework to classify the weight of evidence using a five-level hierarchy (i.e., "causal relationship";
"likely to be causal relationship"; "suggestive of, but not sufficient to infer, a causal relationship";
"inadequate to infer the presence or absence of a causal relationship"; or "not likely to be a causal
relationship" as described in Table II of the Preamble (U.S. EPA. 2015).
This framework for making causality determinations was recently reviewed by an ad hoc
committee of the National Academies of Sciences, Engineering, and Medicine. The committee broadly
endorsed the framework, concluding that it "allows EPA to draw conclusions that integrate scientific
findings across multiple study designs and disciplines, as required by the [Clean Air Act]" (NASEM.
2022). The committee further provided recommendations on approaches to increase transparency in how
evidence is integrated and on other aspects of the ISA causality framework. EPA is currently evaluating
the committee's recommendations and anticipates incorporating appropriate changes to the framework in
future ISAs and documenting these changes in a future revision of the Preamble.
IS.2 Pb Source to Concentration
This section characterizes the current state of atmospheric and environmental science relevant to
understanding Pb exposure and Pb-related health and ecological effects described in subsequent sections.
It builds on previous research reviewed in the 2013 Pb ISA (U.S. EPA. 2013a) and previous Pb AQCDs
(U.S. EPA. 2006. 1986c. 1977). and it emphasizes relevant advances in sources and emissions, fate and
transport, sampling and analysis methods, and concentration observations discussed in greater detail in
Appendix 1 (Lead Source to Concentration). The scope is not limited to airborne Pb from contemporary
emission sources because non-atmospheric processes as well as legacy sources are also relevant for
understanding the effects of air-related Pb. For example, transport and transformation processes in soil
and water after deposition are also relevant. Therefore, current research in other media is also included to
promote understanding of air-related Pb in the context of non-atmospheric sources and media.
In previous ISAs, an up-to-date review of air emissions, monitoring, and concentration trends has
been accomplished through a combination of analysis of U.S. EPA monitoring network data and a
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synthesis of observations reported in the peer-reviewed literature. Reference data such as estimates of
total emissions, coverage of network monitors, average concentrations, and concentration trends can
become out of date before the document is published because these data are so frequently updated. To
facilitate provision of the most current emissions and concentration data from the Pb monitoring network,
a set of relevant maps and graphics that have been routinely included in previous ISAs are now contained
in a separate document titled "Overview of Lead (Pb) Air Quality in the United States" (U.S. EPA.
2022b). Appendix 1 of the Pb ISA provides a literature-based synthesis of recent research on Pb sources,
fate and transport, measurement, and ambient air concentrations.
Section IS.2.1 provides an overview of sources and emissions of Pb in ambient air and other
environmental media. Section IS.2.2 gives descriptions of the fate and transport of Pb in air, soil, and
aqueous media. Section IS.2.3 describes advances in Pb measurement methods, and Section IS.2.4
describes ambient air Pb concentrations, including spatial and temporal variability and the size
distributions of Pb-bearing particulate matter (PM).
15.2.1 Sources and Emissions
Total estimated national Pb emissions to ambient air from the 2020 National Emissions Inventory
(NEI) were 621 tons, with 69% from emissions associated with use of leaded aviation gasoline, 18% from
industrial sources, 9% from fuel combustion, and 3% from wildland fires. All other sources combined
were estimated to account for about 2% of total U.S. Pb emissions estimated by the NEI. Pb emissions
from residential wood combustion are not included in the 2020 NEI but can also be a source in areas
affected by wood smoke in the winter (Appendix 1.2.3). In addition to contemporary Pb emissions into
the atmosphere, historical sources of Pb that are not included in the NEI can potentially contribute to
airborne Pb under some circumstances through the processes of suspension and resuspension
(Appendix 1.3.4). Details of recent research and results of individual studies of Pb emissions from
aviation, industrial sources, stationary fuel combustion, wildfires, automobile traffic and roads, volcanoes,
and legacy sources in the United States are presented in Appendix 1.2.
15.2.2 Fate and Transport
Pb emitted into the atmosphere can be distributed into soil, water, and other media. Pb is mainly
emitted into the air in particulate form. The fate and transport of Pb emitted into the air depends on
particle size, which in turn depends largely on the source. For example, Pb emitted by aircraft using
leaded aviation gas is mainly associated with ultrafine particles smaller than 0.1 |im diameter, while a
large fraction of airborne Pb produced by resuspension of contaminated soil near current or historic
sources can be associated with coarse particles, including particles larger than 10 (mi. Pb-containing
particles are subject to the same atmospheric processes that transport and remove other forms of PM.
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Particle-bound Pb associated with fine PM is transported long distances and found in remote areas, while
Pb associated with coarse PM is more likely to deposit closer to its source. As discussed in
Appendix 1.3.1. the dry deposition rate of particles increases with increasing particle size, effectively
reducing transport distance and atmospheric lifetime. However, depending on the chemical counter-ion,
Pb compounds vary in water solubility, determining the degree to which Pb is removed by wet deposition.
After deposition, resuspension of soil-bound Pb can contribute to airborne concentrations near major Pb
sources (Appendix 1.3.4). There has been little recent research on transport of airborne Pb, beyond a few
individual studies outside the United States that showed agreement of Pb biomonitoring data with
dispersion modeling estimates and chemical transformations of Pb to a more soluble form in polluted air
under specific circumstances (Appendix 1.3.1.2).
In general, fine particulate Pb is mostly soluble and removed from the atmosphere by wet
deposition, and coarse particulate Pb is mostly insoluble and removed from the atmosphere by dry
deposition. Other factors also influence Pb deposition, however. The pH of precipitation can play a role
because Pb solubility increases with decreasing pH. Precipitation can also scavenge insoluble particulate
Pb as an aqueous suspension. Several U.S. studies, some of which have been published since the 2013 Pb
ISA, have reported substantially greater deposition rates in areas near industrial sources than in
nonindustrial areas. Recent studies have also filled in some details about the Pb deposition process,
including studies that indicated Pb deposition increased with elevation and that Pb is enriched in
atmospheric ice nuclei (Appendix 1.3.1.3).
Once deposited in soil, Pb is strongly retained in soil organic material with subsequent Pb fate
and transport through the soil column influenced by several physicochemical factors, including storage in
leaf litter, the amount and decomposition rates of organic matter (OM), composition of organic and
inorganic soil constituents, mobile colloid abundance and composition, microbial activity, and soil pH.
These physicochemical properties are based on soil forming factors: climate, organisms, parent material,
relief (shape of the landscape), time, and anthropogenic input. Soils that differ in these factors will
subsequently have different physicochemical properties and different trends in Pb transport. In general,
leaf litter, low rates of OM decomposition, neutral pH, and soil constituents rich in charged surfaces such
as OM, Fe and Mn oxides, and clay minerals will lead to increased Pb retention and sorption. Conversely,
thin organic layers, increased OM decomposition, acidic pH, increases in anthropogenic Pb, and less
reactive soil constituents such as quartz increase Pb leaching from soils.
In water, runoff from urban or historically industrial areas contains higher Pb concentrations than
runoff from nonurban areas. Recent studies have improved our understanding of relationships between Pb
runoff and street length and density, population density, and land cover, and expanded on the influence of
seasonality and precipitation events on runoff as well as transport and sedimentation. While Pb deposition
has decreased in the last half-century with the phaseout of leaded gasoline and stricter regulation,
accumulated Pb-contaminated sediments in areas with a history of industry and urbanization are
vulnerable to resuspension in water and both down and upstream movement following a disturbance
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event. Dam removal or other disturbances to water bodies can lead to resuspension in water and
dissolution of Pb-contaminated sediment that was previously deposited. With the predicted increase in
future frequency of drought alongside less frequent but more severe precipitation patterns across most of
the United States, the potential for remobilization of such legacy Pb in waterbodies is an area for
consideration.
Additional media besides air, water, and soil are useful for understanding how Pb moves and
changes overtime in the urban environment. It is potentially useful to consider urban soil, resuspended
dust, road dust, and house dust as urban compartments between which Pb can be transported or cycled.
High Pb concentrations are characteristic of urban soil in comparison with other soils and are often related
to legacy sources. Studies in several U.S. cities have explored the high spatial variability of urban soil Pb
concentrations, with hot spots related to income and racial disparities. In recent studies, associations
between airborne Pb and elemental indicators of airborne soil have been observed, suggesting the
potential for contaminated soil to be a source of airborne Pb locally in urban and industrial areas under
some circumstances. Suspension of urban soil into the air can also be a source of Pb in house dust
(Appendix 1.3.4).
IS.2.3 Sampling and Analysis
There are two Federal Reference Methods (FRMs) for sample collection of airborne Pb. The
FRM for Pb in total suspended particulate (Pb-TSP) requires a high-volume sampler and is required for
all source-oriented NAAQS surveillance monitors. The FRM for Pb associated with PMio (Pb-PMio) is
acceptable for Pb NAAQS surveillance monitoring at locations where the expected Pb concentration does
not approach the NAAQS and in the absence of nearby sources of Pb associated with particles greater
than 10 |im diameter. Variability in high-volume TSP sampler collection efficiency associated with
effects of wind speed and sampler orientation for particles larger than 10 (mi has been a serious concern
since the sampler was first implemented for TSP and Pb-TSP sampling. Recent research confirmed that
sampling effectiveness decreased with particle size for coarse particles and varied with wind speed and
sampler orientation. A number of alternative manufacturer-designated low-volume TSP samplers have
been developed, but recent studies showed that their sampling effectiveness also decreases with particle
size for coarse particles. The Pb-PMio FRM is not as vulnerable to sampling errors associated with the
Pb-TSP FRM because it is based on a strictly defined performance standard, but Pb associated with
particles larger than 10 |im in diameter can be an important contributor to airborne Pb exposure. Other
recent advances in ambient air Pb sampling and analysis included the development of a new Pb analysis
FRM based on inductively coupled plasma mass spectrometry, development of more relevant reference
materials for ambient air Pb sampling and analysis, and development of higher time resolution sampling
and analytical methods.
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IS.2.4
Ambient Air Pb Concentrations
Figure IS-1 is a national map of maximum rolling 3-month average Pb concentrations in counties
with Pb-TSP monitors during the period 2020-2022 (Appendix 1.5.1). Concentrations exceeded
0.15 (ig/m3 in Stark County OH (0.40 (ig/m3), Arecibo PR (0.35 (ig/m3), Pike County AL (0.22 (.ig/rn3).
and Lake County IN (0.16 (.ig/rn3). Several recent studies indicated substantial spatial variability in urban
ambient air Pb concentrations influenced by proximity to local sources or industrial activities. Across
urban and neighborhood scales, these variations in ambient air Pb concentrations may not be captured by
national monitoring networks. Seasonal trends were reported in numerous recent studies, but results were
mixed, and no consistent national pattern was apparent. Size distribution data from samples collected near
roads, near industrial sources, in rural locations, and in urban locations within the United States and the
European Union suggest that Pb size distributions in ambient air have shifted in the 1980s from size
distributions with a mass median diameter usually smaller than 2.5 |im to those with a primary mode
between 2.5-10 |im. No recent studies specifically investigated background Pb concentrations, but a
plausible range of 0.2 to 1 ng/m3 was proposed based on earlier studies in the 2013 Pb ISA (U.S. EPA.
2013a).
• 0.06- 0.10 ug/mA3 (11 sites) O 0.16- 0.20 ug/mA3 (1 site)
Source: (U.S. EPA. 2023).
Figure IS-1 Pb maximum rolling 3-month average in jjg/m3 for the 2020-2022
period.
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IS.3 Trends
Total Pb emissions have steadily decreased for decades, largely due to the elimination of leaded
gasoline use in automobiles before 1996, and in later years because of reductions in emissions from
metals processing sources (U.S. EPA. 2022b. 2013a. 2006). From 1990 to 2020, there has been a steep
decline in total U.S. Pb emissions from about 5 kton/year to less than 1 kton/year (U.S. EPA. 2021). In
some cases, there have been more recent periods of continued decline corresponding to reductions in Pb
emissions from local and regional industrial sources. A quantitative description of the trend in ambient air
concentrations based on monitoring network data is problematic for two reasons. First, air Pb
concentration reporting requirements changed in 2010 from measured Pb concentration at standard
temperature and pressure to Pb concentration measured under local conditions. As a result, daily
concentration, and design value data from before 2010 are not directly comparable to data from after
2010. Second, as numerous monitors have been discontinued because of declining Pb concentrations, the
proportion of monitors located near sources has increased. Pb monitoring network data show that the
national median of maximum 3-month average Pb concentrations across monitoring sites declined by
89% from 1990 to 2010 for a mix of 74 source-oriented and non-source-oriented monitors that operated
continuously through this period (Appendix 1.5.1). For a smaller population of 37 monitors with a higher
proportion of source-oriented monitors that operated continuously from 2010 to 2021, the national
median of maximum 3-month average Pb concentrations across monitoring sites decreased by 88% over
that period (Appendix 1.5.1). This recent decrease was driven by the 2008 NAAQS revision and the
steepest declines were observed over the period from 2012 to 2015 when emissions from sources near
these monitors were being reduced to meet the new 2008 Pb NAAQS requirements (Appendix 1.5.1). The
declining trend since 2010 is therefore more representative of a small number of communities near major
sources than an urban or national median. A national trend is more difficult to assess because the number
of non-source-oriented monitors is small, and their observed concentrations are close to method detection
limits on most days (Appendix 1.5.1). Detailed maps and graphics of changing ambient air Pb
concentrations over time are available in U.S. EPA (2022b).
Changes in the patterns of Pb emissions overtime and between regions of the United States are
also detectable in non-air environmental media and biota. Pb may be retained in soils, sediments, the
shells of long-lived bivalves, or trees, where it provides a historical record of deposition such as phaseout
of Pb from on-road gasoline and reductions in industrial releases. However, information on Pb
atmospheric trends can be difficult to interpret due to the influence of other anthropogenic inputs of Pb
and heterogeneity associated with natural environments. The number of studies that examine trends in Pb
concentration in non-air media at national and regional scales is limited.
Concentrations of Pb in soils (Appendix 11.2.2.1) vary across the United States due to a variety
of natural and anthropogenic factors, including historical Pb deposition. The United States Geological
Survey North American Soil Geochemical Landscapes Project (NASGLP) provides the most
comprehensive and rigorous information on the distribution of Pb across the conterminous United States
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(Smith et al.. 2013). In the NASGLP survey, soil samples were collected from multiple depths at 4,857
sites. In areas with historic depositional input of Pb, the concentration of Pb observed in upper-horizon
soils was often higher than that observed in the bedrock. Figure IS-2C shows the ratio of A-horizon (the
uppermost mineral soil) to C-horizon (a deeper soil sample generally of partially weathered parent
material) Pb concentrations mapped in Woodruff et al. (2015). using inverse-distance weighting methods
derived from the NASGLP survey (Smith et al.. 2013). This map displays areas with increased
concentrations of Pb in A-horizon soils relative to lower horizons, hinting at the lasting effect of
depositional Pb pollution, where historical Pb deposition may have a relatively higher effect on people
and ecosystems. Patterns of elevated A- to C-horizon soil Pb concentrations in Figure IS-2C are
conspicuous in areas with historical anthropogenic sources of Pb. This pattern is observed in the
northeastern United States, with a historically high population density and intensity of industrial
development. Likewise, mapping highlights former Pb smelting and mining sites, for instance in areas
near smelters in Everett and Tacoma, Washington or the Doe Run smelter in Herculaneum, Missouri (the
last Pb smelter in the United States, which closed in 2013). Areas near mining sites, including near
Leadville, Colorado, Cooke City, Montana, and northern Utah, also have a high ratio of A- to C-horizon
Pb. Woodruff et al. (2015) emphasized that no known natural geological process would otherwise explain
elevated A-horizon soils relative to the underlying layers.
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ION ^
A. Lead (Pb) - A Horizon
C. Lead (Pb) - Ratio of A Horizon/C Horizon
Source: Woodruff et al. (2015).
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te * J:
EXPLANATION
PBRC n'ftftg
631.0
211
B. Lead (Pb) - C Horizon
I 'Ml - EE 934
¦ "id! - 10UB
21 -100
0-20
D. Population density (per sq. mile) by county
Figure IS-2 Maps of Pb sampled from (A) A-horizon and (B) C-horizon soils,
(C) the ratio of Pb observed in A-horizon to C-horizon soils, and
(D) population density.
In a regional survey of forest floor soils limited to the northeastern United States featuring
sequential sampling in 1980, 1990, 2002, and 2011, mean soil Pb concentrations decreased from 151 ± 29
(standard error [SE]) mg Pb/kg in 1980 to 68 ± 13 (SE) mg Pb/kg in 2011 and were estimated to decline
2.0 ± 0.3 % per year (Richardson et al.. 2015; Richardson et al.. 2014). A 2019 survey of peri-urban soil
Pb in several southern California counties is illustrative of the regional variability in U.S. soil Pb
concentrations. Soil Pb in the study (mean of 23.9 ±13.8 mg Pb/kg) was elevated relative to the
southwestern U.S. region, but lower than concentrations found at contaminated sites near point sources of
Pb ( ackowiak et al.. 2021). These recent national and regional surveys of soil Pb document the spatial
and temporal patterns of residual pollution resulting from decades of Pb emissions. In general, areas with
higher population density and intensity of industrial activity have higher soil Pb concentrations relative to
rural areas. Recent results from more local studies of individual cities and neighborhoods are consistent
with these results (Appendix 1.3.4).
Quantification of Pb in tree rings can be used to reconstruct histoncal trends of Pb in air pollution
(U.S. EPA. 2006. 1986b. 1977); however, radial transport of Pb, which may vary among species, can
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occur within the tree, contributing uncertainty and reducing precision of such reconstructions.
Additionally, there may be a 10- to 15-year delay in tree ring Pb compared with air concentrations as Pb
deposition leaches through the soil and is absorbed by the tree (U.S. EPA. 2013a). Although trends varied
across tree species and regions in several North American studies published since the 2013 Pb ISA
(Appendix 11.2.2.2). studies identified a temporal pattern of Pb that increased after 1850-1900 and, in
some cases, peaked in 1970-1985, then decreased afterward. Tree ring studies with temporal patterns that
deviate from this pattern were conducted near active industrial point sources of Pb pollution.
Temporal trends of Pb deposition in sediment show distinct peaks associated with leaded gasoline
usage in the United States. These peaks are found globally, corresponding to the specific phaseout periods
for the contributing countries (Appendix 1.3.3.4). Patterns of increasing Pb concentration occurring from
the mid-19th century through the mid-20th century due to early industry as well as agriculture,
weathering, and mining operations are identifiable in North American lake and reservoir sediments.
Following the peak deposition period in the 1960s due to leaded gasoline in North America, widespread
decreases in Pb concentration in sediments are seen over the following half-century, but concentration
values are still higher than pre-19th century levels showing continued deposition, nonpoint
contamination, and/or legacy Pb runoff contributions.
In freshwater environments, no recent studies were identified that examined spatial or temporal
trends in Pb concentration in fish or invertebrates from locations across the United States.
Appendix 11.3.2 summarizes several historical studies reviewed in earlier AQCDs or the 2013 Pb ISA.
Limited evidence from regional studies of temporal trends in freshwater aquatic ecosystems published
since the 2013 Pb ISA, including one of dissolved Pb in Appalachian streams and another of peat cores in
northern Alberta, Canada, suggests that modern atmospheric deposition of Pb is not a major contributor to
Pb concentrations in streams in remote locations (Appendix 11.3.2).
In long-term biomonitoring studies of saltwater biota (Appendix 11.4.2). there is some evidence
of declining Pb concentrations, particularly in studies that began sampling before the 1990s. However,
other studies provide mixed results, with some observations of insignificant change or even increases in
Pb concentrations, likely due to non-air anthropogenic sources. The National Oceanic and Atmospheric
Administration (NOAA) Mussel Watch program has monitored pollutant trends since 1986 via periodic
sampling of bivalve tissue (Mytilus spp. and Crassostrea virginica oysters) and sediment along the U.S.
coastline (Kimbrough et al.. 2008). In general, the highest concentrations of Pb are in bivalves in the
vicinity of urban and industrial areas. Metals concentrations in Mytilus ccdifornianiis were sampled at
long-term biomonitoring sites off the coast of California from 1977 to 2010 (specific years vary by site)
(Melwani et al.. 2014). Decreasing trends were observed at some sites while others showed no significant
trend. In addition to tissues, quantification of chemical variation of elements taken up and deposited in
shells of marine organisms (sclerochronology) provides a temporal record of Pb deposition inputs to
coastal environments. Several studies in bivalves collected off the coast of the eastern United States,
where Pb sources include atmospheric transport by easterly winds, show elevated Pb in shell
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corresponding to the peak of Pb gasoline use in the United States and then declines after that time
(Krausc-Nchring et al.. 2012; Gillikin et al.. 2005). In a synthesis of data from 15 studies from different
geographic locations that quantified Pb in marine bivalves, shell concentrations tended to be higher in
areas near sources of Pb pollution (Cariou et al.. 2017). In addition to bivalves, heavy metals quantified in
horseshoe crab (Limiilus polyphemns) eggs collected along Delaware Bay in 2012 showed a decline in Pb
overtime in a comparison with compiled data from earlier surveys conducted between 1993 and 2000
(Burger and Tsipoura. 2014). In contrast, a decade-long biomonitoring study of metals in the muscle
tissue of dolphinfish (('oryphaena hippiinis) in the southern Gulf of California from 2006-2015 found no
temporal trend in Pb concentrations (Gil-Manriquc et al.. 2022).
Overall, evidence from surveys of Pb in environmental media and biota reflects a decline in
anthropogenic emissions of Pb. However, Pb pollution persists in environmental media and is still
observed in measurable concentrations within biota, particularly near sources of Pb pollution both
historical and current. Long-term monitoring of Pb concentration trends in biota (e.g., the NOAA Mussel
Watch program) and soil surveys covering large spatial extents (e.g., NASGLP) provide essential records
of Pb concentrations in the environment observed across decades and regions.
IS.4 Human Exposure to Ambient Pb
Human exposure to Pb derives from the multiple sources of Pb in the environment and the
various media through which it passes (Appendix 2.1). Air-related pathways of Pb exposure are the focus
of this assessment. However, exposure studies containing Pb concentrations in other media (soil, dietary
sources, consumer products, occupational sources, and ammunition) were included because cumulative
body burden can occur as a result of contributions from multiple exposure pathways (i.e., ingestion of Pb-
containing soil by children) and most Pb biomarker studies do not indicate species or isotopic signature,
making it a challenge to link Pb exposures to specific sources. Air-related Pb exposure pathways include
inhalation of Pb in ambient air along with inhalation and ingestion of Pb in indoor dust and/or outdoor
soil that originated from recent or historic ambient air (e.g., air Pb that has penetrated into the residence
either via the air or tracking of soil), ingestion of Pb in drinking water drawn from surface water
contaminated from atmospheric deposition or contaminated from surface runoff of deposited Pb, and
ingestion of Pb in dietary sources after uptake by plants or livestock of Pb that originated from the
atmosphere. Soil can act as a reservoir for deposited Pb emissions. Exposure to soil contaminated with
deposited Pb can occur through inhalation of resuspended soil as well as ingestion via hand-to-mouth
contact. The primary contribution of ambient air Pb to young children's blood Pb concentrations is
generally due to ingestion of Pb following its deposition to soils and dusts (Appendix 2.1.3.2).
Nonambient air-related exposures include hand-to-mouth contact with dust or chips of peeling Pb-
containing paint or ingestion of Pb in drinking water leached by corroding pipes. Several studies indicate
that Pb-containing paint in the home (or home age used as a surrogate for the presence of Pb paint) are
important residential factors that increase risk of elevated blood Pb (Appendix 2.1.3.2).
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The size distribution of soil or dust particles containing Pb differs from the size distribution of
inhalable ambient Pb-bearing PM (Appendix 2.1.3.1). Airborne particles containing Pb tend to be small
(much of the distribution <10 |im) compared with soil or dust particles containing Pb (-50 |im to several
hundred |im). The size of particles containing Pb that someone may be exposed to can vary due to source
type and proximity to those sources. Ingestion through hand-to-mouth contact is the predominant
exposure pathway for the larger particles in soil and dust containing Pb.
A number of monitoring and modeling techniques have been employed for estimating Pb
exposures and associated BLLs. Environmental Pb concentration data can be collected from ambient air
Pb monitors, soil Pb samples, dust Pb samples, and dietary Pb samples to estimate human exposure.
Exposure estimation error depends, in part, on the collection efficiency of these methods. Models, such as
the Integrated Exposure Uptake Biokinetic (IEUBK) model, coupling of the Stochastic Human Exposure
and Dose Simulation (SHEDS) and IEUBK models (SHEDS-IEUBK), and the All-Ages Lead Model,
simulate human exposure to Pb from multiple sources and through intake routes of inhalation and
ingestion. Children's exposure to Pb is modeled using inputs including soil Pb concentration, air Pb
concentration, dietary Pb intake including drinking water and Pb-dust ingestion, human activity, and
biokinetic factors. The relative contribution from specific exposure pathways (e.g., water, diet, soil,
ambient air) to blood Pb concentrations is situation specific. Measurements and/or assumptions can be
utilized when formulating the model inputs; errors in measurements and assumptions have the potential to
propagate through exposure models. Biomarkers, such as blood Pb, can also be used to provide
information about exposure (Appendix 2.3).
IS.5 Toxicokinetics
The majority of Pb in the body is found in bone (roughly 90% in adults, 70% in children); only
about 1% of Pb is found in the blood. Pb in blood is primarily (-99%) bound to red blood cells (RBCs). It
has been suggested that the small fraction of Pb in plasma (<1%) may be the more biologically labile and
toxicologically active fraction of the circulating Pb. The relationship between Pb in blood and plasma is
approximately linear at relatively low daily Pb intakes (i.e., <10 |ig/kg per day) and at blood Pb
concentrations <25 (ig/dL and becomes curvilinear at higher blood Pb concentrations due to saturable
binding to RBC proteins. As BLL increases and the higher affinity binding sites for Pb in RBCs become
saturated, a larger fraction of the blood Pb is available in plasma to distribute to brain and other tissues.
See Appendix 2.2.2 for additional details.
The half-life of Pb in blood is approximately 20-30 days in adults and ahalf-life of
approximately 6 days has been estimated based on data for children under the age of 3 years. An abrupt
change in Pb uptake gives rise to a relatively rapid change in blood Pb, with a new quasi steady-state
achieved in approximately 75-100 days (i.e., 3-4 times the blood elimination half-life). A slower phase of
Pb clearance from the blood may become evident with longer observation periods following a decrease in
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exposure due to the gradual redistribution of Pb among bone and other compartments. See
Appendix 2.3.5 for additional details. Absorbed Pb is excreted primarily in urine and feces, with sweat,
saliva, hair, nails, and breast milk being minor routes of excretion. Approximately 30% of intravenously
injected Pb in humans (40%-50% in beagles and baboons) is excreted via urine and feces during the first
20 days following administration (Leggett. 1993). The kinetics of urinary excretion following a single
dose of Pb is similar to that of blood (Chamberlain et al.. 1978). likely because Pb in urine derives largely
from Pb in plasma. See Appendix 2.2.3 for additional details.
The burden of Pb in the body may be viewed as divided between a dominant slow compartment
(bone) and smaller fast compartment(s) (soft tissues). Pb uptake to and elimination from soft tissues is
much faster than in bone. Pb accumulates in bone regions undergoing the most active calcification at the
time of exposure. Pb accumulation is thought to occur predominantly in cortical bone during childhood
and in both cortical and trabecular bone in adulthood. However, several considerations complicate the
dichotomy between Pb accumulation in trabecular versus cortical bone. For example, the tibia is generally
considered a cortical bone with less than 1% trabecular bone at its midshaft but is 55%-75% trabecular
bone toward the ends of the bone. A high bone formation rate in early childhood results in the rapid
uptake of circulating Pb into mineralizing bone; however, in early childhood, bone Pb is also recycled to
other tissue compartments or excreted in accordance with a high bone resorption rate. Thus, much of the
Pb acquired early in life is not permanently fixed in the bone due to rapid bone formation and
reabsorption. See Appendix 2.2.2.2 for additional details.
The exchange of Pb from plasma to the bone surface is a relatively rapid process. Pb in bone
becomes distributed in trabecular bone and the denser cortical bone. The proportion of cortical to
trabecular bone in the human body varies by age, but on average is about 80% cortical to 20% trabecular.
Of the bone types, trabecular bone is more reflective of recent exposures than is cortical bone because of
the slower turnover rate and lower blood perfusion of cortical bone. Some Pb diffuses to kinetically
deeper bone regions where it is relatively inert, particularly in adults. These bone compartments are much
more labile in infants and children than in adults as reflected by half-times for movement of Pb from bone
into the plasma (e.g., cortical half-time = 0.23 years at birth, 3.7 years at 15 years of age, and 23 years in
adults; trabecular half-time = 0.23 years at birth, 2.0 years at 15 years of age, and 3.8 years in adults)
(Leggett. 1993). See Appendix 2.3.5 for additional details.
Evidence for maternal-to-fetal transfer of Pb in humans is derived from umbilical cord blood to
maternal blood Pb ratios (i.e., cord blood Pb concentration divided by mother's blood Pb concentration).
Group mean ratios range from about 0.7 to 1.0 at the time of delivery for mean maternal BLLs ranging
from 1.7 to 8.6 (ig/dL. Transplacental transfer of Pb may be facilitated by an increase in the plasma/blood
Pb concentration ratio during pregnancy. Maternal-to-fetal transfer of Pb appears to be related partly to
the mobilization of Pb from the maternal skeleton. See Appendix 2.2.2.4 for additional details.
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IS.6 Pb Biomarkers
Overall, BLLs have been decreasing among U.S. children and adults over the past 45 years. The
geometric mean BLL for the entire U.S. population was 0.753 (ig/dL (95% CI: 0.723, 0.784), based on the
2017-2018 NHANES data (CDC. 2021). Among children aged 1-5 years, the geometric mean was
slightly lower, at 0.670 (ig/dL (95% CI: 0.600, 0.748). By comparison, the 1976-1980 NHANES showed
a geometric mean blood Pb of 15.2 (ig/dL (95% CI: 14.3, 16.1) in children aged 1-5 years. In addition,
the gap in BLLs between non-Hispanic Black children and children of different racial/ethnic groups, aged
1-5 and 6-10 years, has decreased overtime, as shown by 1999-2000 to 2015-2016 NHANES data. See
Appendix 2.4.1 for additional details.
Blood Pb is dependent on both the recent exposure history of the individual and the long-term
exposure history, which determines body burden and Pb in bone. The contribution of bone Pb to blood Pb
varies depending on the duration and intensity of the exposure, age, and various other physiological
stressors (e.g., nutritional status, pregnancy, menopause, extended bed rest, hyperparathyroidism) that
may affect bone remodeling, which occurs continuously under normal circumstances. In children, blood
Pb is both an index of recent exposure and potentially an index of body burden, largely due to faster
exchange of Pb to and from bone than in adults. In adults and children whose exposure to Pb has
effectively ceased or greatly decreased, there is a rapid decline in blood Pb over the first few months
followed by a more gradual, slow decline in blood Pb concentrations over the period of years due to the
gradual release of Pb from bone. Bone Pb is an index of cumulative exposure and body burden. Bone
compartments should be recognized as reflective of differing exposure periods, with Pb in trabecular bone
exchanging with the blood more rapidly than Pb in cortical bone. Consequently, Pb in cortical bone is a
better marker of cumulative exposure, while Pb in trabecular bone is more likely to be correlated with
blood Pb, even in adults. See Appendix 2.2.2 and 2.3.5 for additional details.
It is important to recognize that from a single measurement of blood Pb, it cannot be determined
the extent to which blood Pb reflects recent exposure, movement of Pb from bone into blood from
historical exposures, or both recent and historical exposures. Additionally, a single measurement of blood
Pb cannot inform whether an individual is at a steady-state blood Pb concentration or whether blood Pb is
changing because of a change in Pb exposure. In contrast, multiple blood Pb concentrations overtime can
provide more insight into cumulative exposures and average Pb body burdens overtime. The degree to
which repeated sampling will reflect the actual long-term time-weighted average blood Pb concentration
depends on the sampling frequency in relation to variability in exposure. High variability in Pb exposures
can produce episodic (or periodic) oscillations in blood Pb concentration that may not be captured with
infrequent samples. Furthermore, similar blood Pb concentrations in two individuals (or populations),
regardless of their age, do not necessarily translate to similar body burdens or similar exposure histories.
The blood Pb measurement method (capillary or venous) may also influence measured blood Pb
concentrations because of a positive bias in capillary sample measurement and contamination of
fingertips where samples were collected. See Appendix 2.3.2 for additional details.
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The concentration of Pb in urine follows blood Pb concentration. There is added complexity with
Pb in urine because concentration is also dependent upon urine flow rate (see Appendix 2.2.3). which
requires timed urine samples that is often not feasible in epidemiologic studies. Other biomarkers have
been utilized to a lesser extent (e.g., Pb in teeth, hair, and saliva) because of complications with
environmental contamination or inconsistent associations with blood Pb. See Appendix 2.3 for additional
details.
IS.7 Evaluation of the Health Effects of Pb
IS.7.1 Connections Among Health Effects
Broad health effect categories organized by organ system are evaluated separately in the
appendices of this ISA, though the mechanisms underlying disease progression may overlap and are not
necessarily restricted to a single organ system. This section provides a brief overview of how the
relationship between Pb exposure and a variety of health outcomes may be related or affect one another.
Pb-induced injuries can take place via complex pathways within the body. The health effects of
Pb can be triggered by both direct and indirect actions within an organ but can also cause systemic
changes that can affect other areas of the body. Pb can directly bind to cellular proteins and in some
instances can displace biologically relevant enzymes leading both to ion imbalance and initiation of
inflammation and oxidative stress. Because the circulatory system is connected to all body systems,
effects of damage in one organ system may contribute to health effects in another. Pb-induced systemic
inflammation and oxidative stress can trigger systemic responses in multiple organs.
There is crosstalk between organ systems in the body. For example, the nervous system regulates
the development and function of many organs and thus, modulation of the nervous system by Pb exposure
(see Appendix 3) can have widespread effects. Pb has been shown to disrupt the network of signaling
between the hypothalamus, pituitary, and adrenal and gonadotropic axes, which have important
implications in the regulation of development, reproduction, cardiovascular function, and respiratory
function. This is of particular concern with Pb exposure early in life as proper organ development requires
proper hormonal and cell signaling cues. Disruption of these processes early in life can lead to lasting
changes in organ structure and function. In a similar manner, the function of the liver, kidneys, and
cardiovascular system are also linked. The liver plays a major role in the generation, trafficking, and
metabolism of fatty acids and cholesterol, which are trafficked throughout the body for use in other
organs. Alterations in cholesterol and fatty acid homeostasis by Pb can affect the organ systems that use
these resources. The renin-angiotensin system provides another means of crosstalk between the kidney,
liver, and cardiovascular systems. Renin, produced in the kidney, processes angiotensin, produced by the
liver, which can promote vascular contraction. Pb-induced increases in angiotensin processing can lead to
various effects including increased vascular constriction and increased blood pressure (BP). Chronic
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increases in BP can lead to kidney damage. Although these examples are not exhaustive, they highlight
means by which Pb-induced effects in one organ could lead to systemic effects capable of eliciting
multiple health effects.
While all systems of the body are connected intrinsically, most of the available research
examining the health effects of Pb exposures focuses on specific health endpoints. Thus, this ISA includes
separate supporting appendices for Nervous System Effects (Appendix 3). Cardiovascular Effects
(Appendix 4). Renal Effects (Appendix 5). Immune System Effects (Appendix 6). Hematological Effects
(Appendix 7). Reproductive and Developmental Effects (Appendix 8). Effects on Other Organ Systems
and Mortality (Appendix 9). and Cancer (Appendix 10).
IS.7.2 Biological Plausibility
Biological plausibility can strengthen the basis for causal inference (U.S. EPA. 2015). In this
ISA, biological plausibility is part of the weight-of-evidence analysis that considers the totality of the
health effects evidence, including consistency and coherence of effects described in experimental and
observational studies. Each of the human health appendices (Appendix 3-Appendix 10) includes a
biological plausibility section that summarizes the evidence for potential pathways by which Pb
exposures could result in adverse health outcomes at the population level. Although there is some overlap
in the potential pathways between the appendices, each biological plausibility section is tailored to the
specific health outcome category for which causality determinations are made.
Each of the biological plausibility sections includes a figure illustrating possible pathways that
connect Pb exposures with health outcomes. Pathways are based on evidence evaluated in previous
assessments, both AQCDs and IS As, as well as evidence from more recent studies. The accompanying
text characterizes the evidence upon which the figures are based, including results of studies
demonstrating specific effects related to Pb exposure and considerations of physiology and
pathophysiology. Together, the figure and text portray the available evidence that supports the biological
plausibility of Pb exposure leading to specific health outcomes. Gaps in the evidence base (e.g., health
endpoints for which studies have not been conducted) are represented by corresponding gaps in the
figures and are identified in the accompanying text.
In the model figure below (Figure IS-3), which serves as an illustrative overview of the biological
plausibility figures in the health appendices, each box represents evidence demonstrated in a study or
group of studies for a particular effect related to Pb exposure. While most of the studies used to develop
the figures are experimental studies (i.e., animal toxicological and in vitro studies), some observational
epidemiologic studies also contribute to the pathways. These epidemiologic studies generally comprise
effects observed at the population level. The boxes are arranged horizontally, with boxes on the left side
representing initial effects that reflect early biological responses and boxes to the right representing
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intermediate (i.e., subclinical or clinical) effects and effects at the population level. The boxes are color
coded according to their position in the exposure to outcome continuum.
Pb
Exposure
r* -1
Intermediate
Effect 2
j i
I
* 1
Intermediate
Effect 3
Note: The boxes above represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and
the arrows indicate a proposed relationship between these effects. Solid arrows, in contrast to dotted arrows, denote evidence of
essentiality as provided, for example, by an inhibitor of the pathway or a genetic knockout model used in an experimental study
involving Pb exposure. Shading around multiple boxes is used to denote a grouping of these effects. Arrows may connect individual
boxes, groupings of boxes, and individual boxes within groupings of boxes. Progression of effects is generally depicted from left to
right and color coded (white, exposure; green, initial effect; blue, intermediate effect; orange, effect at the population level or a key
clinical effect). Here, population-level effects generally reflect results of epidemiologic studies. When there are gaps in the evidence
base, there are complementary gaps in the figure and the accompanying text below.
Figure IS-3 Illustrative figure for potential biological pathways for health
effects following Pb exposure.
The arrows that connect the boxes indicate a progression of effects resulting from exposure to Pb.
In most cases, arrows are dotted (arrow 1), denoting a possible relationship between the effects. While
most arrows point from left to right, some arrows point from right to left, reflecting progression of effects
in the opposite direction or a feedback loop (arrow 2). In a few cases, the arrows are solid (arrow 2),
indicating that progression from ilthe upstream to downstream effect has been shown to occur as a direct
result of Pb exposure. This relationship between the boxes, where the upstream effect is necessary for
progression to the downstream effect, is termed essentiality (OECD. 2018). Evidence supporting
essentiality is generally provided by experimental studies using pharmacologic agents (i.e., inhibitors) or
animal models that are genetic knockouts. The use of solid lines, as opposed to dotted lines, reflects the
availability of specific experimental evidence that Pb exposure results in an upstream effect which is
necessary for progression to a downstream effect.
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In the figures, upstream effects are sometimes linked to multiple downstream effects. To illustrate
this proposed relationship using a minimum number of arrows, downstream effects are grouped together
within a larger shaded box and a single arrow (arrow 3) connecting the upstream effect represented by a
single box to the outside of the downstream shaded box containing the multiple effects. Multiple
upstream effects may similarly be linked to a single downstream effect using an arrow (arrow 4) that
originates from the outside of a shaded box, which contains multiple effects, to an individual downstream
box. In addition, arrows sometimes connect one individual upstream effect to an individual downstream
effect that is contained within a larger shaded box (arrow 2) or two individual effects both contained
within separate larger shaded boxes (arrow 5). Thus, arrows may connect individual boxes, groupings of
boxes, and individual boxes within groupings of boxes depending on the proposed relationships between
effects represented by the boxes.
IS.7.3 Summary of Health Effects Evidence
Results from health studies, supported by the evidence from atmospheric chemistry and exposure
assessment studies, contribute to the causality determinations made for the health outcomes evaluated in
this ISA. Recent evidence is considered in combination with the evidence presented in the 2013 Pb ISA.
This ISA evaluates the available health effects evidence and presents causality determinations for 30 health
effect categories. In addition to updated causality determinations for the various health outcomes that were
evaluated in the 2013 Pb ISA, this ISA includes three new causality determinations for social cognition and
behavior in children, metabolic effects, and total (nonaccidental) mortality. The causality determinations
from this ISA and their relation to the conclusions from the 2013 Pb ISA are summarized in Table IS-1.
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Table IS-1 Summary of causality determinations by health outcome
Outcome Group
Health Outcome
Causality Determination
Cognitive effects
Causal
Externalizing behaviors: attention,
impulsivity, and hyperactivity
Causal
Nervous System Effects
Ascertained During Childhood,
Adolescent, and Young Adult
Lifestages
Externalizing behaviors: conduct disorders,
aggression, and criminal behavior
Likely to be causal
Internalizing behaviors: anxiety and
depression
Likely to be causal
Motor function
Likely to be causal
Sensory function
^Suggestive
Social cognition and behavior
+Suggestive
Cognitive effects
^Causal
Nervous System Effects
Psychopathological effects
Likely to be causal
Ascertained During Adult
Lifestages
Sensory function
Suggestive
Neurodegenerative disease
^Suggestive
Cardiovascular Effects3
Cardiovascular effects and cardiovascular-
related mortality
Causal
Renal Effects
Renal effects
^Causal
Immunosuppression
Likely to be causal
Immune System Effects'5
Sensitization and allergic response
^Suggestive
Autoimmunity and autoimmune disease
Inadequate
Hematological Effects
Hematological effects, including altered
heme synthesis and decreased RBC
survival and function
Causal
Pregnancy and birth outcomes
^Likely to be causal
Reproductive and Developmental
Development
Causal
Effects
Female reproductive function
^Likely to be causal
Male reproductive function
Causal
Hepatic effects
^Suggestive
Metabolic effects
+ lnadequate
Gastrointestinal effects
Inadequate
Effects on Other Organ Systems
and Mortality
Endocrine system effects
Inadequate
Musculoskeletal effects
Likely to be causal
Ocular health effects
Inadequate
Respiratory effects
Inadequate
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Outcome Group
Health Outcome
Causality Determination
Total (nonaccidental) mortality Total (nonaccidental) mortality +Causalc
Cancer Cancer Likely to be causal
RBC = red blood cell.
+Denotes new causality determination.
| or | Denotes change in causality determination from 2013 Pb ISA.
aThe 2013 Pb ISA made four causality determinations with respect to cardiovascular disease (CVD), including BP and
hypertension (causal), subclinical atherosclerosis (suggestive), coronary heart disease (CHD; causaf), and cerebrovascular
disease (inadequate). This ISA follows the precedent set by the 2019 Particulate Matter and 2020 Ozone ISAs (U.S. EPA. 2020.
2019) by making a single causality determination for cardiovascular effects.
bThe evidence for immune system effects in this ISA is organized based on the World Health Organization's Guidance for
Immunotoxicity Risk Assessment for Chemicals (IPCS. 2012). For comparison with the causality determinations issued in the 2013
Pb ISA, the evidence considered for "sensitization and allergic response" maps closely with "atopic and inflammatory disease," the
"immunosuppression" section largely overlaps with "decreased host resistance," and the evaluation of "autoimmunity and
autoimmune disease" includes consideration of the same endpoints as "autoimmunity."
The 2013 Pb ISA evaluated studies of all-cause mortality together with studies examining cardiovascular mortality and did not
issue a separate causality determination for total mortality.
There is substantial evidence across scientific disciplines (i.e., animal toxicology and
epidemiology) demonstrating that Pb exposure can result in a range of health effects, including nervous
system effects in children and adults, cardiovascular effects, and reproductive and developmental effects.
The evidence that supports these causality determinations includes studies examining the potential
biological pathways that provide evidence of biological plausibility; studies examining the broader health
effects evidence spanning scientific disciplines; and studies examining issues related to exposure
assessment, toxicokinetics, and biomarkers of Pb exposure. The subsequent sections focus on health
outcome categories for which the health effects evidence indicates a "causal relationship" or a "likely to
be causal relationship" and outcome categories for which a previous "causal relationship" or a "likely to
be causal relationship" has been changed (i.e., "likely to be causal" changed to "suggestive of, but not
sufficient to infer a causal relationship"). The evidence for Pb exposure and health effects that is
"suggestive of, but not sufficient to infer, a causal relationship" or "inadequate to infer the presence or
absence of a causal relationship" is noted in Table IS-1 and discussed more fully in the respective health
effects appendices (Appendix 3-Appendix 10).
IS.7.3.1 Nervous System Effects Ascertained During Childhood, Adolescent, and
Young Adult Lifestages
While Pb affects nearly every organ system, the nervous system appears to be one of the most
sensitive targets. The collective body of recent epidemiologic and toxicological evidence, along with
evidence detailed in the 2013 Pb ISA (U.S. EPA. 2013a). demonstrates effects of Pb exposure on a range
of nervous system effects ascertained during childhood, adolescent, and young adult lifestages. These
effects include cognitive function (Appendix 3.5.1). externalizing behaviors (Appendix 3.5.2 and
Appendix 3.5.3). internalizing behaviors (Appendix 3.5.4). and motor function (Appendix 3.5.5). Tables
at the end of each of the ensuing subsections provide a summary of the evidence from epidemiologic and
animal toxicological studies, highlighting the state of the science in the 2013 Pb ISA and summarizing the
recent evidence (Table IS-2A through Table IS-2F).
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IS.7.3.1.1 Cognitive Function in Children
The epidemiologic and toxicological evidence evaluated in the 2013 Pb ISA was sufficient to
conclude that there is a "causal relationship" between Pb exposure and decrements in cognitive function
in children. The strongest evidence supporting this determination came from multiple prospective studies
conducted in diverse populations that consistently reported associations between higher blood and tooth
Pb levels and lower full-scale IQ (FSIQ), executive function, and academic performance and
achievement. Most studies examined representative populations and had moderate to high follow-up
participation without indication of selective participation among children with higher BLLs and lower
cognitive function (i.e., no evidence of selection bias). Associations between BLL and cognitive function
decrements were found with adjustment for several potential confounding factors, most commonly
socioeconomic status (SES), parental IQ, parental education, and parental caregiving quality. In children
ages 4-11 years, associations were found with prenatal, early childhood, childhood average, and
concurrent BLLs in populations with mean or group BLLs in the range of 2-8 (ig/dL. At the time of the
previous review, neither epidemiologic nor experimental animal evidence had identified an individual
critical lifestage or duration of Pb exposure within childhood associated with cognitive function
decrements. Several epidemiologic studies observed a supralinear C-R relationship (i.e., larger decrement
in cognitive function per unit increase in blood Pb level in children in the lower range of the study
population blood Pb distribution). Additionally, a threshold for cognitive function decrements was not
discernible from the available evidence (i.e., examination of early childhood blood Pb or concurrent blood
Pb in the range of <1 to 10 (.ig/dL). Epidemiologic evidence in children was coherent with animal
toxicological studies that observed consistent evidence of Pb-induced impairments in learning, memory,
and executive function in juvenile animals. Several studies in animals indicated learning impairments
with prenatal, lactational, postlactational and lifetime (with or without prenatal) Pb exposures that
resulted in BLLs of 10-25 (ig/dL. The biological plausibility for Pb-associated cognitive function
decrements was supported by observations of Pb-induced impairments in neurogenesis, synaptogenesis,
synaptic pruning, long-term potentiation, and neurotransmitter function in the hippocampus, prefrontal
cortex, and nucleus accumbens.
Recent studies support the conclusion from the 2013 Pb ISA that Pb-associated cognitive effects
in children occur in populations with mean BLLs between 2 and 8 (ig/dL (Appendix 3.5.1.6.1). This
conclusion continues to be based on studies that examined early childhood BLLs (i.e., age <3 years),
considered peak BLLs in their analysis (i.e., peak <10 (.ig/dL). or examined concurrent BLLs in young
children aged 4 years. Some recent studies reported associations of Pb exposure with cognitive effects
among children with mean BLLs <2 (ig/dL; however, those studies do not have the aforementioned
attributes and there is heterogeneity in both the magnitude and direction of the associations at the lowest
blood Pb concentrations. The observed heterogeneity may be explained in part by the distribution of at-
risk factors among the populations studied, including sex, maternal stress, and co-exposures to other
metals and neurotoxic chemicals. Additionally, the available studies do not generally have the sensitivity
(Cooper et al.. 2016) to detect the effect or hazard at these very low BLLs. Therefore, the heterogeneity
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observed in studies with low mean BLLs (i.e., < 2 (ig/dL) does not weaken the larger body of evidence
supporting the association of Pb exposure with cognitive effects in children at BLLs <5 (ig/dL. The
collective body of epidemiologic studies provides no evidence of a threshold for cognitive effects in
children across the range of BLLs examined. Epidemiologic and toxicological studies also continue to
strongly support the finding that Pb exposure during multiple lifestages (prenatal through
adolescence/early adulthood) is associated with cognitive function decrements in children and young
adults. Recent toxicological studies extend the evidence indicating that early-life exposures are associated
with cognitive effects that persisted later into adolescence and adulthood. Biological plausibility is
provided by studies that describe pathways involving the interaction of Pb with cellular proteins, in some
cases competing with and displacing other biologically relevant cations, leading to increased oxidative
stress and the presence of inflammation, which can have widespread impacts on brain structure and
function, as well as disruptions of calcium ion (Ca2+) signaling that can result in alteration in brain
signaling and contribute to the development of neurological impairments.
Given consistency of the results from epidemiologic studies of FSIQ, Bayley Mental
Development Index (MDI), and academic performance and achievement, as well as the coherence of
evidence across epidemiologic and animal toxicological studies of learning and memory, the overall
evidence remains sufficient to conclude that there is a causal relationship between Pb exposure and
cognitive effects in children.
Table IS-2A Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and nervous system effects ascertained
during childhood, adolescent, and young adult lifestages
Cognitive Effects in Children: Causal Relationship (IS.7.3.1.1 and Appendix 3.5.1)
Evidence from the 2013 Pb ISA
Clear evidence of cognitive function decrements (as
measured by FSIQ, academic performance, and executive
function) was reported in young children (4 to 11 yr old) with
mean or group BLLs measured at various lifestages and
time periods between 2 and 8 |jg/dL. Clear support from
animal toxicological studies that demonstrate decrements in
learning, memory, and executive function with dietary
exposures.
Evidence from the 2024 Pb ISA
Recent longitudinal epidemiologic studies with group
or population means <5 |jg/dL add to the evidence,
generally supporting conclusions from the 2013 Pb
ISA. Heterogeneity in the magnitude and direction of
the associations with FSIQ, which was potentially
explained by modeling choices or modification of the
association by exposure to other metals, sex, or
maternal stress, does not weaken inference from the
large body of supporting evidence. Recent
experimental animal studies provide consistent
evidence that Pb exposure results in learning and
memory impairments, with developmental periods
potentially representing a more sensitive window for
exposure.
BLL = blood lead level; FSIQ = full-scale intelligence quotient; ISA = Integrated Science Assessment; Pb = lead; yr = year(s).
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IS.7.3.1.2 Externalizing Behaviors: Attention, Impulsivity, and Hyperactivity in Children
The evidence presented in the 2013 Pb ISA was sufficient to conclude that there is "a causal
relationship" between Pb exposure and effects on attention, impulsivity, and hyperactivity in children.
Several prospective studies demonstrated associations between blood or tooth Pb levels measured years
before outcomes with attention decrements and hyperactivity in children 7-20 years old, as assessed using
objective neuropsychological tests and/or parent and teacher ratings, which are generally reliable and
valid instruments that predict functionally important outcomes. Most of these prospective studies
examined representative populations without indication of selection bias. The results from prospective
studies were adjusted for potential confounding by SES as well as parental education and caregiving
quality, with some studies also considering parental cognitive function, birth outcomes, substance abuse,
and nutritional factors. BLLs were associated with attention decrements and hyperactivity in populations
with prenatal (maternal or cord), age 3-60-month average, age 6 year, or lifetime average (to age 11-
13 years) mean BLLs of 7 to 14 (ig/dL, and groups with age 30-month BLLs >10 (ig/dL. Most well-
conducted cross-sectional studies that adjusted for potential confounding factors supported these findings,
noting associations of attention decrements, impulsivity, and hyperactivity in children ages 5-7.5 years
with concurrent BLLs with means of 5-5.4 (ig/dL. There were a small number of studies of diagnosed
attention-deficit/hyperactivity disorder (ADHD), which were limited by cross-sectional or case-control
study designs, inconsistent adjustment for SES and parental education, and lack of consideration for
potential confounding by parental caregiving quality. Animal toxicological studies reported increases in
impulsivity or impaired response inhibition in animals with postweaning and lifetime Pb exposures that
resulted in BLLs of 11 to 30 (ig/dL. There was biological plausibility for Pb-associated attention
decrements, impulsivity, and hyperactivity provided by observations of Pb-induced alterations in
neurogenesis, synaptic pruning, and dopamine transmission in the prefrontal cerebral cortex, cerebellum,
and hippocampus.
The largest uncertainty addressed by the recent evidence base is the previous lack of prospective
studies examining ADHD (Appendix 3.5.2.4-3.5.2.5). The bulk of the recent evidence comprises
prospective studies that establish the temporality of the association between Pb exposure and parent or
teacher ratings of ADHD symptoms and clinical ADHD. Across studies, associations were observed with
tooth Pb concentrations, childhood BLLs (<6 (.ig/dL). and with maternal or cord BLLs (2-5 (ig/dL).
Studies of caregiver-reported ADHD symptoms generally report associations of BLLs with composite
indices, but there is some support to indicate that the associations with impulsivity and hyperactivity
symptoms are stronger than the associations with inattention symptoms. Some studies addressed the
validity of caregiver-assessed outcomes by evaluating internal consistency, and one study addressed
reliability/validity concerns by using structural equation modeling to create latent factors for inattention
and hyperactivity-impulsivity for each informant. Rating scales used in these studies are generally reliable
and valid instruments that predict functionally important outcomes. Confounder adjustment has also
become more consistent across recent studies of ADHD. Another recent prospective epidemiologic study
examined clinical ADHD diagnoses after adjusting for parental education and SES, although not quality
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of parental caregiving. In this study, children with BLLs between 5 and 10 (ig/dL (measured <4 years old)
had increased odds of clinically diagnosed ADHD at approximately 6 years of age compared with
children with BLLs <2 (ig/dL. Additionally, a small number of recent studies also serve to extend the
lower bound of the mean BLLs that were observed to be associated with attention, impulsivity, and
hyperactivity in the 2013 Pb ISA. These prospective studies with mean maternal and cord BLLs <5 (ig/dL
report associations with some measures of inattention and impulsivity (Appendix 3.5.2.1-3.5.2.3). Across
studies, there is uncertainty regarding the patterns of exposure that are associated with maternal and cord
BLLs and BLLs in older children, because they may be influenced by higher past exposures.
In summary, the coherence of evidence across epidemiologic and toxicological studies of
externalizing behaviors, as well as biological plausibility provided by studies that outline pathways by
which Pb may interfere with the normal development of externalizing behaviors is sufficient to
conclude that there is a causal relationship between Pb exposure and attention, impulsivity, and
hyperactivity.
Table IS-2B
Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and nervous system effects ascertained
during childhood, adolescent, and young adult lifestages
Externalizing Behaviors: Attention, Impulsivity, and Hyperactivity in Children:
Causal Relationship (IS.7.3.1.2 and Appendix 3.5.2)
Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Clear evidence of attention decrements, impulsivity,
and hyperactivity (assessed using objective
neuropsychological tests and parent and teacher
ratings) was observed in children 7-20 yr. The
strongest evidence for blood Pb-associated
increases in these behaviors was found in
prospective studies examining prenatal (maternal
or cord), age 3-60 mo, age 6 yr, or lifetime average
(to age 11-13 yr) mean BLLs of 7 to 14 |jg/dL and
groups with early childhood (age 30 mo) BLLs
>10 |jg/dL. Biological plausibility was provided by
animal toxicological studies demonstrating
impulsivity or impaired response inhibition with
relevant prenatal, lactational, postrotational, and
lifetime Pb exposures.
A small number of recent studies of children with population or
group mean BLLs <5 |jg/dL contribute to the body of evidence,
supporting and extending conclusions from the 2013 Pb ISA.
The majority of recent studies rely on parent and teacher
ratings of ADHD symptoms; notably, confounder adjustment
remained inconsistent across these studies. However,
prospective studies of ADHD, including a study of clinical
ADHD that controlled for parental education and SES,
although not quality of parental caregiving reported positive
associations. Findings from studies of rodents and nonhuman
primates indicate that Pb exposure changes behavior in ways
consistent with increased impulsivity while experimental animal
studies of hyperactivity remain inconsistent, potentially due to
differential exposure and testing windows (hyperactivity was
consistently observed with lactational exposure).
ADHD = attention-deficit/hyperactivity disorder; BLL = blood lead level; ISA = Integrated Science Assessment; mo = month(s);
Pb = lead; SES = socioeconomic status; yr = year(s).
IS.7.3.1.3 Externalizing Behaviors: Conduct Disorders, Aggression, and Criminal Behavior in
Children, Adolescents, and Young Adults
The 2013 Pb ISA concluded that "a causal relationship is likely to exist" between Pb exposure
and conduct disorders in children and young adults. This determination was based on several prospective
cohort studies that consistently indicated that higher earlier childhood (e.g., age 30 months, 6 years) or
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lifetime average (to age 11-13 years) BLLs or tooth (from ages 6-8 years, generally reflecting prenatal
and early childhood Pb exposure) Pb levels are associated with criminal offenses in children and young
adults ages 19-24 years and with higher parent and teacher ratings of behaviors related to conduct
disorders in children ages 7-17 years. Positive associations between Pb exposure and conduct disorders
were found in populations with mean BLLs of 7-14 (.ig/dL. These associations were found without
indication of strong selection bias and with adjustment for SES, parental education and IQ, parental
caregiving quality, family functioning, smoking, and substance abuse. Associations in populations with
lower BLLs that are not influenced by higher earlier Pb exposures were not well characterized.
Toxicological evidence for Pb-induced aggression in animals is inconsistent, with increases in aggression
found in some studies of adult animals with gestational and lifetime Pb exposure, but not juvenile
animals.
Recent epidemiologic studies support and extend the findings from the 2013 Pb ISA
(Appendix 3.5.3.1). The strongest evidence comes from recent prospective cohort studies of 1) self-
reported conduct and aggression-related outcomes, and 2) external measures of delinquency
(e.g., criminal arrests, school suspensions). These studies evaluated outcomes among individuals ages 7-
33 years in relation to earlier (or cumulative) blood and bone Pb levels. Mean and/or median BLLs ranged
from 2.3 to 8.7 (ig/dL (measured 6.5 to 13 years) in studies reporting positive associations with self-
reported conduct and aggression-related outcomes, though were higher in studies reporting positive
associations with external measures of delinquency (mean: 14.4 (ig/dL; measured prenatal to 6 years).
There are no recent animal toxicological studies at BLLs relevant to humans; thus, the central uncertainty
present in the 2013 Pb ISA database remains: there is limited and inconsistent supporting evidence from
animal toxicological studies. Despite the lack of recent studies examining aggression in animals exposed
to Pb, Pb-induced changes on many neurochemical endpoints that contribute to aggressive behaviors have
been reported in experimental animal studies (Appendix 3.3). which provides biological plausibility for
Pb-related conduct disorders and aggression.
Given the consistent positive associations observed across various populations and based on
multiple outcome assessment approaches at relevant Pb exposure levels, there is sufficient evidence to
conclude that there is likely to be a causal relationship between Pb exposure and conduct disorders,
aggression, and criminal behavior.
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Table IS-2C Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and nervous system effects ascertained
during childhood, adolescent, and young adult lifestages
Externalizing Behaviors: Conduct Disorders, Aggression, and Criminal Behavior:
Likely to Be Causal (IS.7.3.1.3 and Appendix 3.5.3)
Evidence from the 2013 Pb ISA
Prospective epidemiologic studies demonstrated that early
childhood (age 30 mo, 6 yr) or lifetime average (to age 11-
13 yr) BLLs or tooth Pb levels (from ages 6-8 yr, generally
reflecting prenatal and early childhood Pb exposure) are
associated with criminal offenses in young adults ages 19-
24 yr and with higher parent and teacher ratings of behaviors
related to conduct disorders in children ages 8-17 yr. Pb-
associated increases in conduct disorders were found in
populations with mean BLLs 7 to 14 |jg/dL; associations with
lower BLLs as observed in cross-sectional studies were
likely to be influenced by higher earlier Pb exposures. There
is coherence in epidemiologic findings among related
measures of conduct disorders. Evidence of Pb-induced
aggression in animals was mixed, with increases in
aggression found in some studies of adult animals with
gestational plus lifetime Pb exposure, but not juvenile
animals. The lack of clear biological plausibility produces
some uncertainty.
Evidence from the 2024 Pb ISA
Several prospective studies add to the evidence,
particularly in providing evidence of positive
associations between Pb exposure and direct
aggressive measures, such as physical violence. A
limited number of recent studies examine crime or
delinquency and generally observed positive
associations. Studies evaluated the associations
among individuals ages 7-33 yr in relation to earlier
(or cumulative) Pb levels. In the studies of self-
reported conduct and aggression-related outcomes,
mean BLLs were 2.3 to 8.7 |jg/dL (ages 6.5 to 13 yr),
while in studies of external measures of delinquency,
they were higher (e.g., mean 14.4 |jg/dL; prenatal to
6 yr). Studies generally controlled for most relevant
confounders. There were no recent experimental
animal studies of aggression; thus, toxicological
evidence remains limited and inconsistent.
BLL = blood lead level; ISA = Integrated Science Assessment; Pb = lead; yr = year(s).
IS.7.3.1.4 Internalizing Behaviors: Anxiety and Depression
The evidence evaluated in the 2013 Pb ISA was sufficient to conclude that "a causal relationship
is likely to exist" between Pb exposure and internalizing behaviors in children. Prospective studies in a
few populations demonstrated associations between increases in early lifetime average blood (mean:
-14 (ig/dL) or childhood tooth (from ages 6-8 years, generally reflecting prenatal and early childhood Pb
exposure) Pb levels with higher parent and teacher ratings of internalizing behaviors, such as withdrawn
behavior and symptoms of depression and anxiety in children ages 8-13 years. The available evidence of
study participation by BLL and parental and teacher ratings do not suggest a high likelihood of selection
bias in these studies. The results from a few cross-sectional studies in populations with mean concurrent
BLLs of 5 (ig/dL were inconsistent. Pb was positively associated with internalizing behaviors in models
that adjusted for maternal education and SES-related variables, though consideration of potential
confounding by parental caregiving quality was inconsistent. Despite some uncertainty in the
epidemiologic evidence regarding potential confounding and inconsistency in the supporting cross-
sectional studies, biological plausibility for the effects of Pb on internalizing behaviors was provided by a
few studies in animals with dietary lactational Pb exposure, with some evidence at BLLs relevant to
humans. Biological plausibility findings included Pb-induced changes in the hypothalamic-pituitary-
adrenal axis and dopaminergic and gamma-aminobutyric-acid (GABA)-ergic systems.
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Several recent longitudinal epidemiologic studies with high to moderate participation rates used
an expanded array of instruments to assess internalizing behaviors and continue to provide support for
associations with BLLs (childhood average, prenatal, and postnatal BLLs <7 (ig/dL; Appendix 3.5.4.1).
The majority of analyses controlled for important potential confounders including the quality of parental
caregiving, which was less frequently considered by studies included in the 2013 Pb ISA. A limited
number of studies aimed to distinguish between the types of internalizing behaviors associated with Pb
exposure and demonstrated stronger support for Pb-associated anxiety compared with depression. Recent
animal toxicological studies are coherent with the epidemiologic evidence and largely support and expand
evidence of increases in anxiety-like behavior in Pb-exposed rodents with peak BLLs ranging from three
to greater than 30 (ig/dL, the lower end of which is lower than evidence from the 2013 Pb ISA.
Overall, given consistent evidence from both recent and previously reviewed prospective
epidemiologic studies, with some remaining uncertainties regarding potential confounding by quality of
parental caregiving, the evidence is sufficient to conclude that there is likely to be a causal
relationship between Pb exposure and internalizing behaviors in children.
Table IS-2D Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and nervous system effects ascertained
during childhood, adolescent, and young adult lifestages
Internalizing Behaviors: Anxiety and Depression: Likely to Be Causal (IS.7.3.1.4 and Appendix 3.5.4)
Evidence from the 2013 Pb ISA
Prospective epidemiologic studies reported associations of
higher lifetime average blood (mean: -14 |jg/dL) or
childhood tooth (from ages 6-8 yr, generally reflecting
prenatal and early childhood Pb exposure) Pb levels with
higher parent and teacher ratings of internalizing behaviors
such as symptoms of depression or anxiety and withdrawn
behavior in children ages 8-13 yr. Consideration of potential
confounding by parental caregiving was not consistent and
findings from cross-sectional studies in populations ages 5
and 7 yrwith mean BLLs of 5 |jg/dL were mixed. Animal
toxicological studies demonstrate depression-like behaviors
and increases in emotionality with relevant lactational
exposures.
Evidence from the 2024 Pb ISA
Recent longitudinal studies report consistent
associations between BLLs and internalizing in
multiple countries with mean blood Pb concentrations
typically <7 |jg/dL (prenatal, early childhood, lifetime
average). Recent studies used parent or teacher
ratings to assess internalizing behaviors, i.e., the
Child Behavior Checklist, Strengths and Difficulties
Questionnaire, Behavioral Assessment System for
Children, and Caregiver-Teacher Report Form.
Increased anxiety-like behaviors in rodents were
demonstrated at lower exposure levels (3-30 |jg/dL)
following developmental Pb exposure.
BLL = blood lead level; ISA = Integrated Science Assessment; mo = month(s); Pb = lead; yr = year(s).
IS.7.3.1.5 Motor Function
The evidence presented in the 2013 Pb ISA was sufficient to conclude that "a causal relationship
is likely to exist" between Pb exposure and decrements in motor function in children. This determination
was based on strong evidence from prospective studies that reported that higher maternal, neonatal,
concurrent, and lifetime average BLLs were associated with lower scores on fine and gross motor
function tests among children ages 4.5-6 years and that higher earlier childhood (ages 0-5-year average;
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age 78 months) BLLs were associated with lower scores on fine and gross motor function tests among
children ages 15-17 years. The means for these blood Pb metrics ranged from 4.8 to 12 (ig/dL. These
studies included adjustment for several potential confounding factors, including SES, parental caregiving
quality, and child health, and did not have indications of substantial selection bias. Evidence from cross-
sectional studies was less consistent. The biological plausibility for associations observed in children is
provided by a study that found poorer balance in male mice with relevant gestational to early postnatal
(postnatal day 10) Pb exposures.
Several recent birth cohort studies observed consistently lower scores on the Bayley Psychomotor
Developmental Index (PDI) in association with higher maternal Pb exposure (no clear pattern by trimester
of pregnancy; means or geometric means: 1.4 to 6.5 (.ig/dL). cord BLL (means: 1.2 to 5.6 (.ig/dL). and
postnatal concurrent BLL (2.85-4.87 (ig/dL). Pb-associated decrements in motor function were also
observed in neonates and in some, but not all studies of toddlers that assessed motor function using the
Gesell scale or children's abilities to perform certain tasks indicative of gross motor function. Evidence
from recent toxicological studies is coherent with the epidemiologic evidence, indicating that
developmental Pb exposure in rodents induces deficits in motor function at mean BLLs <30 (ig/dL. These
new studies illustrate effects of Pb exposure across a range of gross and fine motor development in novel
paradigms. In addition to the effect on rotarod performance described in the 2013 Pb ISA, recent studies
observed Pb-induced decrements in righting reflex, negative geotaxis reflex, ascending wire mesh, and
forelimb hang tests. The available studies demonstrate consistent effects on motor function, but given the
disparate effects examined do not provide evidence of consistent results for any specific test. Overall, the
evidence is sufficient to conclude that there is likely to be a causal relationship between Pb exposure
and motor function in children.
Table IS-2E Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and nervous system effects ascertained
during childhood, adolescent, and young adult lifestages
Motor Function: Likely to Be Causal (IS.7.3.1.5 and Appendix 3.5.5)
Evidence from the 2013 Pb ISA
Prospective epidemiologic studies provided evidence of
associations of fine and gross motor function decrements in
children ages 4-17 yr with lifetime average BLLs and with
BLLs measured at various time periods with means
generally ranging from 4.8 to 12 |jg/dL. Results were
inconsistent in cross-sectional studies with concurrent BLL
means 2-5 |jg/dL. Limited evidence in animal toxicological
studies with relevant Pb exposures.
Evidence from the 2024 Pb ISA
Several recent birth cohort studies report lower
scores on the Bayley PDI at ages 12 to 36 mo in
association with higher maternal Pb exposure, cord
BLL, and postnatal concurrent BLL. Limited biological
plausibility is provided by a small number of recent
toxicological studies showing various effects on motor
function in rodent models with developmental Pb
exposure resulting in BLLs <30 |jg/dL.
BLL = blood lead level; ISA = Integrated Science Assessment; mo = month(s); Pb = lead; PDI = Psychomotor Developmental
Index; yr = year(s).
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IS.7.3.1.6 Sensory Organ Function
The 2013 Pb ISA presented two causality determinations relating to sensory function in children:
auditory function and visual function. The evidence was sufficient to conclude that "a causal relationship
is likely to exist" between Pb exposure and auditory function decrements in children, while the evidence
was "inadequate to determine that a causal relationship exists" between Pb exposure and visual function
in children. In this ISA, recent studies inform a single causality determination for sensory organ function.
A prospective epidemiologic study, as well as a few cross-sectional studies, reported associations between
BLLs and hearing loss and auditory processing deficits with BLLs measured at various time periods,
including prenatal maternal, neonatal (10 day, mean 4.8 |ig/dL). lifetime average (to age 5 years), and
concurrent (ages 4-19 years; median 8 (ig/dL). Evidence for Pb-associated increases in hearing thresholds
or latencies of auditory evoked potentials was found in adult monkeys with lifetime dietary Pb exposure.
However, these effects in adult animals were found with higher peak or concurrent BLLs (i.e., 33-
150 (ig/dL); thus, the biological plausibility for epidemiologic observations is unclear. Studies examining
visual effects were of limited quantity and relevance.
Recent cross-sectional and case-control studies have continued to demonstrate associations
between BLLs and hearing loss in young children (aged 3-7, BLLs ~3 to 6 (ig/dL) and adolescents (aged
12-19, BLLs ~1 to 8 (.ig/dL). particularly at higher frequencies. This is coherent with previously noted
evidence in adult monkeys. Recent experimental animal studies have not further evaluated hearing
thresholds in nonhuman primates at more relevant BLLs, but a recent study reported an 8-12 dB upward
shift in brainstem auditory evoked potentials (BAEPs) between 4 and 32 kHz in young adult mice
exposed during adolescence (peak BLLs 29 (.ig/dL). Similar studies did not detect differences in BAEP in
rodents with lower peak BLLs (3 to 8 (.ig/dL): however, decrements in auditory processing (e.g., sound
discrimination and localization) were demonstrated at these lower BLLs (8 (ig/dL). A few recent
epidemiologic studies also evaluated BAEP with inconsistent results. Evidence of visual function remains
limited and inconsistent.
In conclusion, recent evidence is generally consistent with the evidence presented in the 2013 Pb
ISA. Cross-sectional and case-control epidemiologic studies provide some support for positive
associations between Pb exposure and impaired hearing/auditory processing. Toxicological evidence for
Pb-induced auditory functioning, particularly in studies with BLLs relevant to humans, remains limited.
Taken together, the evidence is suggestive of, but not sufficient to infer, a causal relationship
between Pb exposure and sensory function in children.4
4The Preamble to the ISA, which was published after the release of the 2013 Pb ISA, included some minor changes
to the weight-of-evidence descriptors for the five-level causality hierarchy . These changes resulted in the evidence
for Pb exposure and effects on sensory organ function being more consistent with examples of evidence that is
"suggestive of, but not sufficient to infer, a causal relationship. " Therefore, the change from "likely to be causal" to
"suggestive of, but not sufficient to infer, a causal relationship " reflects minor changes to the causal framework,
rather than a weakening of the evidence base. See Appendix 3.5.6.4 of the Nervous System Effects Appendix for
further discussion.
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Table IS-2F Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and nervous system effects ascertained
during childhood, adolescent, and young adult lifestages
Sensory Function: Suggestive of, but Not Sufficient to Infer, a Causal Relationship
(IS.7.3.1.6 and Appendix 3.5.6)
Evidence from the 2013 Pb ISA
A prospective epidemiologic study and large cross-sectional
studies indicated associations between BLLs and increased
hearing thresholds at ages 4-19 yr. Across studies,
associations were found with BLLs measured at various time
periods, including prenatal maternal, neonatal (10 d, mean
4.8 |jg/dL), lifetime average, and concurrent (ages 4-19 yr)
BLLs (median 8 |jg/dL). The lack of biological plausibility in
animals with relevant exposures produces some uncertainty.
The available epidemiologic and toxicological evidence for
visual function is of insufficient quantity, quality, and
consistency.
Evidence from the 2024 Pb ISA
Recent cross-sectional and case-control studies have
continued to demonstrate positive associations
between BLLs and hearing loss in young children
(aged 3-7, BLLs ~3 to 6 |jg/dL) and adolescents
(aged 12-19, BLLs ~1 to 8 |jg/dL). Experimental
animal studies evaluating young adult rodents found
increases in BAEP thresholds at mean BLLs of
29 |jg/dL but not at lower mean BLLs (3 to 8 pg/dL);
however, mice with mean peak BLLs of 8 |jg/dL had
deficits in auditory processing, The epidemiologic and
toxicological evidence for visual function was not
extended.
BAEP = brainstem auditory evoked potentials; BLL = blood lead level; d = day(s); ISA = Integrated Science Assessment;
Pb = lead; yr = year(s).
IS.7.3.2 Nervous System Effects Ascertained During Adult Lifestages
This ISA presents causality determinations for four nervous systems outcomes ascertained during
adult lifestages, including cognitive function, psychopathological effects, sensory organ function, and
neurodegenerative diseases. The available evidence is "suggestive of, but not sufficient to infer, a causal
relationship" between Pb exposure and: (1) sensory organ function (Appendix 3.6.3) and (2)
neurodegenerative disease (Appendix 3.6.4). Evidence related to these outcomes is discussed in the
Nervous System Effects Appendix. Cognitive effects (IS.7.3.2.1) and psychopathological effects
(IS.7.3.2.2), for which evidence supports a causal relationship and a likely to be causal relationship with
Pb exposure, respectively, are discussed in more detail in the ensuing sections. Table IS-3A and
Table IS-3B provides a summary of the evidence from epidemiologic and animal toxicological studies
related to these outcomes, highlighting the recent evidence in comparison with the evidence available in
the 2013 Pb ISA.
IS.7.3.2.1 Cognitive Function in Adults
The 2013 Pb ISA concluded that "a causal relationship is likely to exist between" long-term
cumulative exposure to Pb and cognitive function decrements in adults. This causality determination was
supported by prospective studies in the Normative Aging Study (NAS) and Baltimore Memory Study
(BMS) cohorts that reported that higher cumulative exposure metrics, including baseline tibia (means 19,
20 |ig/g) or patella (mean 25 |ig/g) Pb levels, were associated with declines in cognitive function in adults
(age >50 years) over 2- to 4-year periods. These associations were noted in models adjusted for a range of
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potential confounding factors, including age, education, SES, current alcohol use, and current smoking.
Supporting evidence was provided by cross-sectional analyses of the NAS, BMS, and the Nurses" Health
Study, which observed inverse associations between cognitive function and Pb exposure that were
stronger (i.e., greater in magnitude) for bone Pb levels compared with concurrent BLLs. Cross-sectional
studies also considered more potential confounding factors, including dietary factors, physical activity,
medication use, and comorbid conditions. The range of exposures and health outcomes examined in many
of these studies reduced the likelihood of participation bias, specifically by adults with higher Pb
exposure and lower cognitive function. The specific timing, frequency, duration, and magnitude of Pb
exposures contributing to the associations observed with bone Pb levels was not discernible from the
evidence. The effects of recent Pb exposures on cognitive function decrements in adults were indicated in
studies of Pb-exposed workers, although these studies did not consider potential confounding by other
workplace exposures. Biological plausibility for the observed associations was provided by animal
toxicological studies demonstrating that relevant lifetime Pb exposures from gestation, birth, or after
weaning induced learning impairments in adult animals, as well as evidence the Pb exposure altered
neurotransmitter function in hippocampus, prefrontal cortex, and nucleus accumbens.
Recent prospective cohort studies with longer follow-up periods, multiple and repeatedly
measured cognitive outcomes, and adjustment for multiple risk factors and confounders reduce
uncertainties and strengthen the overall evidence related to the association of Pb exposure with cognitive
function in adulthood. Specifically, recent cohort studies indicate that higher adult bone Pb levels (tibia
mean range: 10.5 to 21.6 (ig/g, patella mean range: 12.6 to 30.6 |ig/g) were associated with poor cognitive
function/performance during young-, mid- or older-adulthood periods (Appendix 3.6.1). A few recent
prospective studies also observed associations between childhood BLLs (mean range: 3.4 (ig/dL to
10.99 (ig/dL at 7-12 years of age) and decrements in IQ and cognitive domains during late adolescence
(18-19 years) and mid-adulthood (38-45 years) after adjustment for demographic and socioeconomic
factors, maternal IQ, and childhood IQ scores. These findings provide new insight into the persistence of
Pb-associated cognitive function decrements. There was some variability in the associations across the
various domains of cognitive function tested within studies; however, higher Pb levels were associated
with decrements in full-scale IQ (verbal comprehension, perceptual reasoning, working memory, and
processing speed IQs), global cognitive function, executive function, visuospatial skills, attention,
learning, and memory. Discordant Pb associations across domains of cognitive function likely reflect
inherent biologic variability or differences in the outcome pathophysiology as opposed to inconsistency in
the evidence. In addition to potential confounders considered in studies evaluated in the 2013 Pb ISA,
recent studies control for additional behavioral, clinical, and neighborhood level factors. Results from
recent toxicological studies are coherent with the epidemiologic evidence and provide evidence that
exposure to Pb during adulthood impairs learning and memory function in rodents with exposure
resulting in mean BLLs <30 (ig/dL (means BLLs: 8-8.8 (ig/dL; peak BLLs: 11-28 (.ig/dL). Additionally, a
few recent studies in juvenile rodents provide some support for the association between Pb exposure
during adolescence and cognitive impairment, but the evidence is less consistent.
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In summary, recent prospective epidemiologic studies address uncertainties from the 2013 Pb
ISA and expand and strengthen the previous body of evidence. Results from recent epidemiologic studies
of childhood Pb exposure and cognitive effects in adults are coherent with previous studies in nonhuman
primates demonstrating cognitive impairment following early-life exposure to Pb and are further
supported by experimental animal studies providing biological plausibility for the observed associations.
Overall, the collective evidence is sufficient to conclude that there is a causal relationship between
Pb exposure and cognitive effects in adults.
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Table IS-3A
Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and nervous system effects ascertained
during adult lifestages
Cognitive Effects in Adults: Likely to Be Causal (IS.7.3.2.1 and Appendix 3.6.1)
Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Prospective studies in the NAS and BMS cohorts indicated
associations of higher baseline tibia (means 19, 20 pg/g) or
patella (mean 25 pg/g) Pb levels with declines in cognitive
function in adults (age >50 yr) over 2- to 4-yr periods. While
the specific covariates differed between studies, these bone
Pb-associated cognitive function decrements were found
with adjustment for potential confounding factors such as
age, education, SES, current alcohol use, and current
smoking. Supporting evidence is provided by cross-sectional
analyses of the NAS, BMS, and the Nurses' Health Study,
which found stronger associations with bone Pb level than
concurrent BLL. Cross-sectional studies also considered
more potential confounding factors, including dietary factors,
physical activity, medication use, and comorbid conditions.
Biological plausibility for the effects of Pb exposure on
cognitive function decrements in adults is provided by
findings that relevant lifetime Pb exposures from gestation,
birth, or after weaning induce learning impairments in adult
animals and by evidence for the effects of Pb altering
neurotransmitter function in hippocampus, prefrontal cortex,
and nucleus accumbens.
Recent longitudinal epidemiologic studies with longer
follow-up periods, multiple and repeatedly measured
cognitive outcomes, and consideration of multiple risk
factors/confounders provide additional evidence of
associations between cumulative and early childhood
exposure to Pb and cognitive decrements in adults.
These studies reduce uncertainties and strengthen
the overall evidence related to the association of Pb
exposure with cognitive function in adulthood. Some
uncertainties related to the frequency, duration, and
magnitude of Pb exposures associated with cognitive
decrements remain. Several recent studies of rodents
with exposure resulting in mean BLLs <30 |jg/dL add
to the evidence informing the association of Pb
exposure with measures of learning and memory in
rodents exposed throughout adulthood.
BMS = Baltimore Memory Study; ISA = Integrated Science Assessment; NAS = Normative Aging Study; Pb = lead;
SES = socioeconomic status.
IS.7.3.2.2 Psychopathological Effects in Adults
The evidence presented in the 2013 Pb ISA was sufficient to conclude that "a causal relationship
is likely to exist" between Pb exposures and psychopathological effects in adults. This causality
determination was based on a small body of epidemiologic evidence that demonstrated consistent positive
associations between concurrent blood or bone Pb levels and self-reported symptoms of depression,
anxiety, and panic disorder in large studies of adults (i.e., NHANES, NAS). Epidemiologic associations
were observed in study populations of young (20-39 years old) and older (44-98 years old) adults.
Because of the cross-sectional design of the epidemiologic studies, there was uncertainty regarding the
temporal sequence between Pb exposure and psychopathological symptoms in adults. This uncertainty is
somewhat reduced with results for tibia Pb because it is an indicator of cumulative Pb exposure. Still,
because these studies included adults with likely higher past Pb exposures, uncertainty exists as to the Pb
exposure level, timing, frequency, and duration contributing to the associations observed with blood or
bone Pb levels. The epidemiologic evidence was supported by coherence in animal toxicological studies
that demonstrated depression-like behavior and emotionality in rodents exposed to dietary lactational Pb
with or without additional postlactational exposure. An uncertainty in the toxicological evidence base was
the limited number of studies that administered exposures resulting in BLLs that are relevant to humans.
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Recent evidence from prospective epidemiologic studies provides further support for positive
associations between Pb exposures and pathological effects, including increased internalizing symptoms.
A strength of the recent evidence is that two of the prospective studies reported a greater likelihood of
internalizing symptoms in association with higher childhood BLLs, providing more information on the
timing of Pb exposure associated with psychopathological effects. Notably, supporting evidence from
recent cross-sectional epidemiologic studies conducted in diverse populations is largely inconsistent. The
epidemiologic evidence is supported by coherence with results from an expanded number of toxicological
studies conducted at BLLs relevant to humans. In addition to recent toxicological studies that continue to
provide strong support for Pb-induced anxiety-like behaviors, and persistence of these behaviors,
following developmental and cumulative exposures, there is some novel evidence for an increase in
anxiety-like behavior following adult exposures. Overall, the collective evidence is sufficient to
conclude that there is likely to be a causal relationship between Pb exposure and psychopathological
effects in adults.
Table IS-3B Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and nervous system effects ascertained
during adult lifestages
Psychopathological Effects in Adults: Likely to Be Causal (IS.7.3.2.2 and Appendix 3.6.2)
Evidence from the 2013 Pb ISA Evidence from the 2024 Pb ISA
Cross-sectional studies in a few populations
demonstrate associations of higher concurrent blood
or tibia Pb levels with self-reported symptoms of
depression and anxiety in adults. Pb-associated
depression and anxiety symptoms among adults were
found with adjustment for age, SES, and in the NAS,
daily alcohol intake. The biological plausibility for
epidemiologic evidence is provided by observations of
depression-like behavior in animals with dietary
lactational Pb exposure.
Recent prospective analyses provide additional support for
a positive association between bone and BLLs and
psychopathological effects in older adults, although results
from cross-sectional studies are inconsistent. Recent
toxicological studies in rodents with developmental
exposure continue to provide evidence of anxiety-like
behaviors. Multiple studies demonstrate the persistence of
these effects into adulthood. Additionally, a few recent
studies in rodents demonstrated effects of adult-only Pb
exposures on anxiety-like behavior after 42-126 d of
exposure (BLLs: 7.1 to 28.4 pg/dL), but not following a 30-d
exposure (BLLs: 6.8 to 8.8 |jg/dL).
BLL = blood lead level; d = day(s); NAS = Normative Aging Study; Pb = lead; SES = socioeconomic status.
IS.7.3.3 Cardiovascular Effects and Cardiovascular-Related Mortality
The 2013 Pb ISA made four causality determinations with respect to cardiovascular disease
(CVD), using the U.S. Surgeon General's Report on Smoking as a guideline to group evidence into health
outcome categories (CDC. 2004). The outcome categories evaluated included BP and hypertension,
subclinical atherosclerosis, coronary heart disease (CHD), and cerebrovascular disease. This ISA follows
the precedent set by the 2019 Particulate Matter and 2020 Ozone IS As (U.S. EPA. 2020. 2019) by making
a single causality determination for cardiovascular effects and cardiovascular-related mortality. This
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approach allows for a more holistic evaluation of interrelated health endpoints (e.g., atherosclerosis,
endothelial dysfunction, and increased BP).
The strongest evidence for cardiovascular effects of Pb exposure in the 2013 Pb ISA came from
studies of BP and hypertension, which supported a causal relationship. Several epidemiologic studies
evaluated in the 2013 Pb ISA (U.S. EPA. 2013a) and previous AQCD documents (U.S. EPA. 2006. 1990)
indicated positive associations between biomarkers of Pb exposure in adults and increases in BP and
hypertension risk (Table IS-4). Previous studies do not identify an apparent threshold below which blood
Pb was not significantly associated with changes in BP, for mean adult BLLs ranging from <2 (ig/dL to
34 (ig/dL. Meta-analyses evaluated in the 2013 Pb ISA underscore the consistency and reproducibility of
the Pb-associated increases in BP and hypertension. However, the studies available at the time
represented populations historically exposed to higher levels of Pb, raising uncertainty regarding the
level, timing, frequency, and duration of Pb exposure contributing to the observed associations. In
addition to epidemiologic evidence, the 2013 Pb ISA described a large body of animal toxicological
studies that provided evidence that long-term Pb exposure (>4 weeks) in experimental animals, resulting
in BLLs less than 10 (ig/dL, could result in the onset of hypertension (after a latency period) in
experimental animals that persists long after the cessation of Pb exposure.
The 2013 Pb ISA also presented a large body of evidence indicating a relationship between Pb
exposure and cardiovascular mortality, which helped support a causal relationship in the 2013 Pb ISA for
CHD. Specifically, prospective epidemiologic studies conducted in a number of locations reported that
biomarkers of Pb exposure were associated with risk of mortality from myocardial infarction (MI),
ischemic heart disease (IHD), and CHD. In addition, epidemiologic studies reviewed in the 2013 Pb ISA
included some evidence of a positive association between exposure to Pb and changes in cardio
electrophysiology (e.g., changes in heart rate variability [HRV] and QT interval) and atherosclerotic
plaque formation. These studies, along with animal toxicological studies demonstrating the production of
oxidative stress species that could inactivate the vasodilator nitric oxide, contribute to the biological
plausibility of Pb-induced cardiovascular morbidity and mortality.
Recent studies greatly expand the evidence base from the 2013 Pb ISA and strengthen support for
the relationship between exposure to Pb and cardiovascular effects in adults. Numerous epidemiologic
studies published since the literature cutoff date for the 2013 Pb ISA reported positive associations
between Pb biomarkers and increases in BP and hypertension risk (Appendix 4.3). Specifically, nationally
representative cross-sectional studies in the United States, Canada, and South Korea observed positive
associations between BLLs and systolic BP and/or diastolic BP in adult populations with mean BLLs
ranging from -1.5 to 3 (ig/dL. Notably, in these studies of adult populations, uncertainty remains
regarding the influence of higher past exposures on the level, timing, frequency, and duration of Pb
exposure contributing to the observed associations. The majority of recent analyses consider a wide range
of confounders including demographics, comorbid conditions, antihypertensive medication use, and other
co-exposures to metals such as cadmium (Cd). In addition, there was also an extensive amount of
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literature that considered effect measure modifiers including, sex, age, and race, among others
(Section IS.7.4). Recent animal toxicological studies are coherent with the epidemiologic evidence of
associations. In several recent studies with exposures resulting in BLLs <30 (ig/dL, animals exposed to Pb
had consistent increases in BP when compared with control treated animals. Combined with results from
the 2013 Pb ISA and AQCDs, there is clear and substantial evidence that exposure to Pb results in
increases in measures of BP.
A number of recent prospective cohort studies, including extended analyses of previous
NHANES cohorts, reported consistent positive associations between BLLs and CVD-related mortality
that are of similar magnitude to results from studies evaluated in the 2013 Pb ISA (Section 4.10). These
more recent studies also reported that associations persisted after accounting for risk factors such as
physical activity, serum cholesterol, and Cd levels in blood or urine. Once again, there were consistent
positive associations between BLLs and mortality in populations with low mean BLLs, but the specific
level, timing, frequency, and duration of Pb exposure contributing to CVD mortality in adult populations
with higher past than recent exposure is not discernible from this evidence. Epidemiologic studies of
mortality are consistent not only with the large amount of evidence for changes in BP and hypertension
described above, but also with evidence of associations between blood or bone Pb levels and other
cardiovascular outcomes. Past and recent analyses of the NAS cohort of older adult men indicate positive
associations between bone Pb levels and incident IHD and prolonged QT interval. Additionally, a series
of recent Korea National Health and Nutrition Examination Survey (KNHANES) studies observed
increased 10-year CVD risk with increasing BLLs. These results are coherent with a toxicological study
evaluated in the 2013 Pb ISA demonstrating increased incidence of arrhythmia, atrioventricular block,
and a prolonged ST segment interval in Pb-exposed animals. In general, animal and in vitro toxicological
evidence provides plausible pathways by which exposure to Pb could lead to serious CVD-related
outcomes such as IHD and MI (Appendix 4.11). A notable pathway includes Pb resulting in oxidative
stress and systemic inflammation that could potentially lead to impaired vascular function, a pro-
atherosclerotic environment, and increases in BP. These effects, in particular atherosclerosis and increases
in BP, can lead to MI or stroke that could result in mortality.
Taken together, the recent evidence supports and extends the evidence base reported in the 2013
Pb ISA. A large number of prospective cohort studies reported consistent associations between body Pb
concentrations and cardiovascular outcomes such as increased BP, hypertension, and cardiovascular
mortality. In particular, studies measuring bone Pb levels provided consistent evidence for associations
between cumulative exposures and chronic health outcomes, such as hypertension, and premature
mortality. This evidence is generally supported by cross-sectional epidemiologic studies and is coherent
with evidence from animal toxicological studies, and further supported by experimental animal and in
vitro studies demonstrating biologically plausible pathways through which exposure to Pb could lead to
these outcomes. Thus, there is sufficient evidence to conclude that there is a causal relationship
between Pb exposure and cardiovascular effects and cardiovascular-related mortality.
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Table IS-4 Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and cardiovascular effects and
cardiovascular-related mortality
Cardiovascular Effects and Cardiovascular-Related Mortality: Causal (IS.7.3.3 and Appendix 4)
Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Hypertension'. Prospective epidemiologic studies
consistently reported associations of blood and bone Pb
levels with hypertension incidence and increased BP. These
findings were consistent across multiple high-quality studies
comprising large and diverse populations. Further support
was provided by multiple cross-sectional analyses. While the
adjustment for specific factors varied by study, the collective
body of epidemiologic evidence included adjustment for
multiple potential key confounding factors. Although
epidemiologic studies in adults observed associations in
populations with relatively low mean concurrent BLLs, the
majority of individuals in these adult populations were likely
to have had higher levels of Pb exposure earlier in life. Thus,
there is uncertainty concerning the specific Pb exposure
level, timing, frequency, and duration contributing to the
associations observed in the epidemiologic studies. A causal
relationship of Pb exposure with hypertension is supported
by evidence from experimental animal studies that
demonstrate effects on BP after long-term Pb exposure
resulting in mean BLLs of 10 |jg/dL or greater.
CHD\ Prospective epidemiologic studies of cohorts of adults
during the period 1976-1994 consistently reported positive
associations between BLLs and risk of CVD mortality,
including Ml and IHD. Several other studies reported
associations between Pb biomarkers and incidence of CHD-
related outcomes, including a prospective analysis reporting
increased incidence of IHD (physician confirmed Ml, angina
pectoris) in association with increasing blood and bone Pb
levels.
Recent studies strengthen support for the relationship
between exposure to Pb and cardiovascular effects in
adults. In particular, the strongest evidence continues
to come from studies demonstrating the effect of Pb
on increases in BP. The majority of recent analyses
examining BP consider a wide range of potential
confounders, including demographics, comorbid
conditions, antihypertensive medication use, and
other co-exposures to metals such as Cd. There is
also an extensive amount of literature that considered
effect measure modifiers, including sex, age, and
race, among others. Recent animal toxicological
studies provide additional evidence that exposure to
Pb resulting in BLLs <30 |jg/dL lead to increases in
measures of BP. In addition to recent evidence on BP
and hypertension, there is substantially more
evidence for cardiovascular-related mortality, as well
as some epidemiologic and toxicological evidence for
effects such as changes in cardiac electrophysiology
(e.g., electrocardiography measures of cardiac
depolarization, repolarization, and HRV), arrythmia,
and markers of atherosclerosis. There continues to
be uncertainty regarding the specific Pb exposure
level, timing, frequency, and duration contributing to
the associations observed in the epidemiologic
studies.
BLL = blood lead level; BP = blood pressure; Cd = cadmium; CHD = coronary heart disease; CVD = cardiovascular disease;
HRV = heart rate variability; IHD = ischemic heart disease; Ml = myocardial infarction; Pb = lead.
IS.7.3.4 Renal Effects
The 2013 Pb ISA concluded that evidence was "suggestive of a causal relationship" between Pb
exposure and renal effects. Recent epidemiologic and toxicological studies extend the body of evidence
presented in the 2013 Pb ISA indicating that Pb exposure is associated with reduced kidney function and
kidney damage (Table IS-5). The causality determination in the 2013 Pb ISA was primarily limited by
uncertainty due to the potential for reverse causality, as kidney damage could lead to increased BLLs
through reduced excretion, rather than increased Pb exposure (e.g., elevated BLLs) being a causative
factor of kidney impairment. A number of recent epidemiologic studies address this uncertainty with
prospective study designs that control for baseline kidney function (Appendix 5.6). These studies
demonstrate associations between biomarkers of Pb exposure and incident markers in kidney function in
adults that are independent of baseline kidney function. Additionally, recent animal toxicological studies
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provide further evidence for Pb-induced kidney damage and dysfunction (e.g., morphological changes in
kidney structure, increased glomerular filtration rate, increased serum and urine creatinine, and increased
blood urea nitrogen), supporting the directionality of effects. Combined, the toxicological and
epidemiologic evidence indicates that reverse causality is highly unlikely to explain the epidemiologic
associations between higher BLLs and decreased kidney function in adults. Toxicological studies also
indicate plausible biological pathways connecting Pb exposure to renal effects, including Pb-induced
oxidative stress and increases in BP (Appendix 5.9). Recent prospective epidemiologic evidence of
associations between BLLs and reduced kidney function in adults are observed at BLLs <5 (ig/dL, and a
number of recent toxicological studies extend the evidence base to include effects in rodents with BLLs
<20 (ig/dL. Despite the evidence for associations at relatively low BLLs in adults, these renal outcomes
were most often examined in adults who have been exposed to higher levels of Pb earlier in life, and
uncertainty remains concerning the Pb exposure level, timing, frequency, and duration contributing to the
observed associations.
Collectively, given the consistent evidence provided by prospective epidemiologic studies that
control for baseline renal function and coherent experimental animal evidence for Pb-induced kidney
damage, there is sufficient evidence to conclude that there is a causal relationship between Pb
exposure and renal effects.
Table IS-5 Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and renal effects
Renal Effects: Causal (IS.7.3.4 and Appendix 5)
Evidence from the 2013 Pb ISA
Longitudinal studies reported Pb-associated decrements in
renal function in populations with mean BLLs of 7 and
9 |jg/dL. However, the contributions of higher past Pb
exposures could not be excluded. Additionally, there was
uncertainty due to potential reverse causality in
epidemiologic studies. Animal toxicological studies provided
clear biological plausibility with evidence for Pb-induced
kidney dysfunction at BLLs greater than 30 |jg/dL; however,
evidence in animals with BLLs <20 |jg/dL was generally not
available.
Evidence from the 2024 Pb ISA
Recent toxicological and prospective epidemiologic
studies support and extend conclusions from the 2013
Pb ISA. Notably, prospective studies with baseline
measures of renal function reduce uncertainty
regarding potential reverse causality, providing
additional evidence of Pb-associated decrements in
renal function in adult populations with mean BLLs
<5 |jg/dL. The contribution of higher past Pb
exposures remains an uncertainty. Recent animal
toxicological studies include evidence for renal effects
observed at concentrations <20 |jg/dL.
BLL = blood lead level; ISA = Integrated Science Assessment; Pb = lead.
IS.7.3.5 Immune System Effects
The 2013 Pb ISA issued causality determinations for the effects of Pb exposure on different
aspects of the immune system including atopic and inflammatory responses, decreased host resistance,
and autoimmunity. The evidence in this ISA is organized based on the World Health Organization's
Guidance for Immunotoxicity Risk Assessment for Chemicals (IPCS. 2012). As proposed in this guidance,
this ISA restructures the available evidence into slightly different outcome groups than those in the 2013
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Pb ISA, which include immunosuppression, sensitization and allergic responses, and autoimmunity and
autoimmune disease. For comparison with the causality determinations issued in the 2013 Pb ISA, the
evidence considered for "sensitization and allergic response" maps closely with "atopic and inflammatory
disease," the "immunosuppression" section largely overlaps with "decreased host resistance," and the
evaluation of "autoimmunity and autoimmune disease" includes consideration of the same endpoints as
"autoimmunity." The recent evidence for autoimmunity and autoimmune disease remains "inadequate to
determine the presence or absence of a causal relationship" (Appendix 6.7). The following sections focus
on the evidence for immunosuppression (IS.7.3.5.1) and sensitization and allergic response (IS.7.3.5.2),
which is also summarized in Table IS-6A and Table IS-6B.
IS.7.3.5.1 Immunosuppression
The 2013 Pb ISA concluded that "a causal relationship is likely to exist" between Pb exposures
and decreased host resistance. This causality determination was based primarily on consistent evidence
that exposure to relevant BLLs suppresses the delayed-type hypersensitivity (DTH) response and
increases bacterial titers and subsequent mortality in rodents. Suppressed DTH response is one of the
most consistently reported immune effects associated with Pb exposure in animals and has been reported
following gestational and postnatal exposures to Pb acetate resulting in BLLs ranging from 6.75 to
>100 (ig/dL in rats, mice, and chickens. A limited number of epidemiologic studies reviewed in the 2013
Pb ISA (U.S. EPA. 2013a) indicated a positive association between BLLs and viral and bacterial
infections in children. None of the studies considered potential confounders, however, and most analyzed
populations with higher BLLs (means >10 (ig/dL). Cross-sectional studies of cell-mediated immunity
reported consistent associations between BLL and lower T cell abundance in children, while results from
other studies on lymphocyte activation, macrophages, neutrophils, and natural killer cells were generally
inconsistent or not sufficiently informative (e.g., cross-sectional study designs with limited or no
consideration of potential confounding, and a lack of information on C-R relationship). Biological
plausibility was provided by a number of studies demonstrating Pb-induced suppression of T helper (Th)l
cytokines production (e.g., interferon-gamma [IFN-y]), and decreased macrophage function, both of
which may lead to decreased DTH response and increased incidence of viral and bacterial infection.
Recent toxicological studies provide additional evidence for immunosuppression, including
decreased serum levels of anti-tetanus toxoid (TT) specific immunoglobulin M (IgM) (but not IgG)
antibodies in iron (Fe)-deficient rats exposed to Pb in drinking water (BLL = 16.1 (.ig/dL). Consistent with
findings reported in the 2013 Pb ISA, recent studies show that Pb exposure suppresses the DTH response
(BLL = 18.48 (ig/dL). Recent epidemiologic studies investigating aspects of immunosuppression include
populations with wider age-ranges and much lower mean and median BLLs than studies evaluated in the
2013 Pb ISA. Recent studies also adjust for a wide range of potential confounders, including extensive
consideration of SES factors. Cross-sectional and case-control studies are coherent with the toxicological
evidence, providing consistent evidence of associations between Pb exposure (mean, median, or
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geometric mean BLLs: 1.4-3.15 (ig/dL) and higher viral and bacterial infection prevalence and lower
antibiotic resistance in children and adults. Notably, epidemiologic studies of viral and bacterial infection
used concurrent blood Pb measures, raising uncertainty regarding the temporal sequence between Pb
exposure and immunosuppression and the level, timing, frequency, and duration of Pb exposures that
contributed to the observed associations. Vaccine antibody response, an endpoint that was not examined
in studies evaluated in the 2013 Pb ISA, was evaluated in a birth cohort study and a few cross-sectional
studies that demonstrate generally consistent evidence of an association between BLLs (mean or median
<2 (ig/dL) and decreased virus-neutralizing antibodies in children. Biological plausibility for the observed
associations is provided by recent and previously evaluated toxicological and epidemiologic studies
demonstrating (1) skewing of T cell populations, promoting Th2 cell formation and cytokine production,
(2) decreased IFN-y production, (3) decrements in macrophage function, (4) production of inflammatory
mediators, and (5) disruption of the microbiome.
Collectively, based on strong evidence from toxicological studies consistently demonstrating that
Pb exposures suppress the DTH response and increase susceptibility to bacterial infection in animals and
supporting evidence from epidemiologic studies demonstrating higher Pb-related susceptibility to viral
and bacterial infection, reduced antibiotic resistance, and reduced vaccine antibodies in children. Overall,
there is sufficient evidence to conclude that there is likely to be a causal relationship between Pb
exposure and immunosuppression.
Table IS-6A
Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and immune system effects
Immunosuppression: Likely to Be Causal (IS.7.3.5.1 and Appendix 6.3)
Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Animal toxicological studies were the primary
contributors to the evidence for Pb-induced
immunosuppression. Several studies in rodents
show that dietary Pb exposure producing relevant
BLLs (7-25 |jg/dL) results in increased
susceptibility to bacterial infection and
suppressed DTH. A few cross-sectional
epidemiologic studies indicated positive
associations between Pb and respiratory
infections, but these studies are limited by a lack
of rigorous methodology or consideration for
potential confounding.
Recent toxicological studies demonstrate the ability of Pb to
alter antibody responses, providing additional evidence for the
immunosuppressive effects of Pb. The relationship between Pb
exposure and immunosuppression is further supported by
recent epidemiologic studies, which expand the quantity and
quality of the observational evidence base evaluated in the
2013 Pb ISA. A mix of recent prospective cohort, case-control,
and cross-sectional studies that include more robust
consideration for potential confounding report associations
between low BLLs (<3.5 |jg/dL) and susceptibility to viral and
bacterial infection, reduced antibiotic resistance, and reduced
vaccine antibodies in children.
BLL = blood lead level; DTH = delayed-type hypersensitivity; ISA = Integrated Science Assessment; Pb = lead.
IS.7.3.5.2 Sensitization and Allergic Response
The 2013 Pb ISA concluded that "a causal relationship is likely to exist" between Pb exposures
and an increase in atopic and inflammatory conditions. This causality determination was supported by a
prospective analysis reporting associations between BLLs and increased asthma incidence in children and
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another longitudinal study that observed a positive association between cord BLLs and immediate-type
allergic responses in children that were detected clinically using skin prick tests. Both studies had small
sample sizes, however, and lacked precision (i.e., had wide 95% CIs), which increases the likelihood of
chance findings. The associations observed in the prospective analyses were supported by a cross-
sectional study of BLL-associated parental-reported asthma in children and population-based cross-
sectional studies in children that reported associations between BLL and elevated serum IgE. Notably,
many of the serum IgE studies had limited adjustment for potential confounders and included population
mean BLLs >10 (ig/dL. The epidemiologic findings were coherent with a large body of toxicological
studies that reported physiological responses in animals consistent with the development of allergic
sensitization, including increased lymph node cell proliferation, increased production of Th2 cytokines
such as interleukin 4 (IL-4), increased total serum IgE antibody levels, and misregulated inflammation.
Recent animal toxicological studies relevant to sensitization and allergic response are limited in
number. The available studies report effects of Pb on production of cytokines relevant to immediate-type
hypersensitivity. However, the utility of these data for hazard identification is limited because changes in
cytokine levels (particularly when measured in blood) can be associated with many different types of
tissues and toxicities and may reflect an immune response to tissue injury but not necessarily an impact
on or impairment of immune function. Recent epidemiologic evidence is inconsistent with studies
evaluated in the 2013 Pb ISA. Specifically, whereas a few small prospective studies reviewed in the 2013
Pb ISA supported the presence of an association between BLLs and incident asthma in children, recent
epidemiologic studies of atopic disease, including prospective cohort studies examining asthma, eczema,
and food allergies were generally consistent in reporting a lack of an association in populations with low
BLLs (mean or median BLLs <2 (ig/dL). Similar to cohort studies evaluated in the 2013 Pb ISA, recent
longitudinal analyses are limited in number and have limited statistical power because of low case
numbers. Among other things, limited statistical power results in the reduced likelihood of detecting a
true effect and a reduced likelihood that an observed result reflects a true effect. Notably, recent cross-
sectional NHANES analyses also reported null associations between children's BLLs and asthma,
eczema, and food allergies in much larger study populations. Additionally, recent studies provide
inconsistent evidence for Pb-associated changes in immunological biomarkers involved in
hypersensitivity and allergic response. Whereas there was coherence between the animal toxicological
and epidemiologic evidence evaluated in the 2013 Pb ISA, the recent epidemiologic studies add
considerable uncertainty to the line of evidence that previously provided support for the "likely to be
causal" determination in the 2013 Pb ISA. Overall, given the strong body of toxicological evidence, but
inconsistent results across epidemiologic studies, the collective evidence is suggestive of, but not
sufficient to infer, a causal relationship between Pb exposure and sensitization and allergic
responses.
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Table IS-6B Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and immune system effects
Sensitization and Allergic Response: Suggestive (IS.7.3.5.2 and Appendix 6.4)
Evidence from the 2013 Pb ISA Evidence from the 2024 Pb ISA
A limited number of prospective studies in a few populations
of children ages 1-5 yr reported associations of asthma and
allergy with BLLs prenatal cord BLLs or BLLs. These studies
had small sample sizes and lacked precision (i.e., had wide
95% CIs). The epidemiologic findings are coherent with a
large body of toxicological studies that reported
physiological responses in animals consistent with the
development of allergic sensitization, including increased
lymph node cell proliferation, increased production of Th2
cytokines such as IL-4, increased total serum IgE antibody
levels, and misregulated inflammation.
Several recent epidemiologic studies of sensitization
and allergic response, including prospective birth
cohorts and cross-sectional studies with mean or
median BLLs <2 |jg/d, provide little evidence of an
association between exposure to Pb and atopic
disease, including asthma, eczema, and food
allergies. Similar to the evidence in the 2013 Pb ISA,
a considerable uncertainty in the evidence base is the
limited number of children with asthma in the cohort
studies evaluated. Recent toxicological evidence for
effects of Pb exposure on biomarkers of allergic
disease is sparse but provides some evidence of Pb-
induced changes in IFN-y, a Th1 cytokine known to
play a role in the resolution of asthma.
BLL = blood lead level; CI = confidence interval; IgE = immunoglobulin E; IL-4 = interleukin 4; ISA = Integrated Science
Assessment; Pb = lead; Th = T helper; yr = year(s).
IS.7.3.6 Hematological Effects
The effects of Pb exposure on RBC function and heme synthesis have been extensively studied
over several decades in both human and animal studies. The 1978 NAAQS for Pb were established to
prevent BLLs in most children from exceeding 30 (ig/dL as such levels were associated with impaired
heme synthesis, evidenced by accumulation of protoporphyrin in erythrocytes (U.S. EPA. 1978). The
2013 Pb ISA issued causality determinations for two hematological outcomes: RBC survival and function
and altered heme synthesis. The evidence for both outcomes was "sufficient to conclude that there is a
causal relationship" with Pb exposure. Given the interconnectedness of the effects of Pb on RBC survival
and function and altered heme synthesis, this assessment presents a single causality determination for the
combination of these outcomes. This approach allows for a more holistic evaluation of interrelated health
endpoints, including a discussion of how all individual lines of evidence contribute to the overall
hematological effects causality determination. The evidence available in the 2013 Pb ISA as well as
evidence from recent studies is discussed in the ensuing subsections and summarized in Table IS-7.
Taken together, there is sufficient evidence to conclude that there is a causal relationship between
Pb exposure and hematological effects, including altered heme synthesis and decreased RBC
survival and function.
IS.7.3.6.1 Red Blood Cell Survival and Function
A strong body of evidence from experimental animal studies reviewed in the 2013 Pb ISA
demonstrated that Pb exposures alter several hematological parameters (e.g., hemoglobin [Hb],
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hematocrit, mean corpuscular volume, mean corpuscular Hb), induce oxidative stress (e.g., alter
antioxidant enzyme activities [superoxide dismutase, catalase, glutathione peroxidase], decrease cellular
glutathione, and increase lipid peroxidation), and increase cytotoxicity in RBC precursor cells in rodents
exposed to various forms of Pb via drinking water and gavage resulting in BLLs <30 (ig/dL. Consistent
results were observed in several additional studies in rodents that did not report BLLs. Results from
epidemiologic studies were coherent with the toxicological evidence, including associations between
BLLs and differences in hematological parameters, higher levels of oxidative stress, altered
hematopoiesis, and higher prevalence of anemia. Notably, the epidemiologic evidence consisted of cross-
sectional studies that were conducted in populations with higher mean Pb exposures (i.e., BLLs
>10 (ig/dL), did not thoroughly consider potential confounders, and lacked rigorous statistical
methodology.
Recent toxicological evidence is limited, but studies continue to support the findings from the last
review. The most consistent evidence comes from studies that report decreased Hb levels in rodents
following Pb exposures (BLLs ranging from 7.5 to 14.7 (ig/dL) (Appendix 7.3.2). Recent epidemiologic
studies expand on the evidence presented in the 2013 Pb ISA and are coherent with the experimental
evidence. Although the recent studies are also cross-sectional, they include populations with much lower
BLL means (<10 (ig/dL) and include more robust adjustment for potential confounding, addressing
important uncertainties from the last review. The most consistent epidemiologic evidence indicates an
association between higher BLLs and lower Hb levels in children (Appendix 7.3.1). which is in line with
the evidence from recent experimental animal studies. While the clinical relevance of small mean
decrements in Hb across exposure quintiles is unclear, a few of the recent epidemiologic studies observed
increases in the odds of prevalent anemia in children associated with increasing quantiles of BLLs.
IS.7.3.6.2 Altered Heme Synthesis
As described in the 2013 Pb ISA, a small but consistent body of studies in adult animals reported
that Pb exposures via drinking water and gavage for 15 days to 9 months (resulting in BLLs <30 (ig/dL)
decreased ALAD and ferrochelatase activities. The relationship between Pb exposure and altered heme
synthesis was further supported by several toxicological studies that observed decreased Hb levels in
laboratory animals exposed to Pb. Decreased Hb levels can be a direct indicator of decreased heme
synthesis. Cross-sectional epidemiologic studies provided supporting evidence that concurrent elevated
BLLs are associated with decreased ALAD and ferrochelatase activities and decreased Hb levels in both
adults and children. However, the majority of these studies are limited by the lack of consideration of
potential confounding. Although there were limitations in the epidemiologic evidence, some studies in
children did control for or consider potential confounding, and effects in adults and children in these
studies are coherent with effects observed in animal toxicological studies.
Recent PECOS-relevant studies are limited in number and focus mainly on Hb levels but continue
to provide support for Pb-related alterations in heme synthesis. Notably, recent epidemiologic studies
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indicate an inverse association between BLLs and Hb levels in children. These studies include more
robust statistical methods, expanded consideration of potential confounders, and populations with much
lower BLLs than the studies included in the previous reviews (mean or median BLLs ranging from 3.04
to 8.38 (ig/dL; Appendix 7.3.1). The recent epidemiologic evidence is coherent with recent toxicological
studies, which observed Hb decrements in Pb-exposed mice with BLLs relevant to humans.
Table IS-7 Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and hematological effects
Hematological Effects, Including Altered Heme Synthesis and Decreased Red Blood Cell Survival and
Function: Causal
(IS.7.3.6.2 and Appendix 7)
Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
RBC Survival and Function'. Experimental animal studies
demonstrate that exposures resulting in BLLs relevant to
humans alter several hematological parameters, increase
measures of oxidative stress, and increase cytotoxicity in
RBC precursor cells. Epidemiologic studies find associations
in both adults and children between BLLs and altered
hematological endpoints, higher measures of oxidative
stress, altered hematopoiesis, and higher prevalence of
anemia. The epidemiologic evidence consisted of cross-
sectional studies that were conducted in populations with
high mean Pb exposures and did not thoroughly consider
potential confounders. Additional support for these findings
was provided by toxicological and epidemiologic studies
demonstrating increased intracellular Ca2+ concentrations,
decreased Ca2+/Mg2+ adenosine triphosphatase activity, and
increased phosphatidylserine exposure, establishing
biologically plausibility for Pb-induced changes in RBC
survival.
RBC Survival and Function'. Recent animal
toxicological studies are limited in number, but
consistent with evidence in the 2013 Pb ISA. The
most consistent evidence comes from studies that
report decreased Hb levels in rodents following Pb
exposures (BLLs of 7.5 to 14.7 |jg/dL). Recent
epidemiologic studies include populations with much
lower BLL means than studies in the 2013 Pb ISA
(3.04 to 8.38 |jg/dL) and more robust adjustment for
potential confounding. The most consistent
epidemiologic evidence indicates associations
between higher BLLs and lower Hb levels and higher
prevalence of anemia in children (birth to 11 yr).
Heme Synthesis: Altered heme synthesis (e.g., decreased
ALAD and ferrochelatase activities, and decreased Hb
levels) was demonstrated by a small, but consistent, body of
epidemiologic and toxicological studies with relevant Pb
exposures. Epidemiologic studies were all cross-sectional
and the majority lacked consideration for potential
confounding. Evidence for altered heme synthesis is also
provided by a large body of toxicological and epidemiologic
studies that report lower Hb concentrations in association
with Pb exposure or BLLs.
Heme Synthesis: Recent epidemiologic studies
indicate an inverse association between BLLs and Hb
levels in children. These studies expanded
consideration of potential confounders and include
populations with lower BLLs (mean or median BLLs
ranging from 3.04 to 8.38 |jg/dL). The recent
epidemiologic evidence is coherent with recent
toxicological studies, which also observed Hb
decrements in Pb-exposed mice with BLLs relevant to
humans.
ALAD = 6-aminolevulinic acid dehydratase; BLL = blood lead level; Ca2+ = calcium ion; Hb = hemoglobin; ISA = Integrated Science
Assessment; Mg2+ = magnesium ion; Pb = lead; RBC = red blood cell; yr = year(s).
IS.7.3.7 Reproductive and Developmental Effects
This ISA organizes the reproductive and developmental effects of Pb exposure into four outcome
categories: effects on pregnancy and birth outcomes, effects on development, effects on female
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reproductive function, and effects on male reproductive function. The collective evidence is sufficient to
conclude that there is "likely to be a causal relationship" between Pb exposure and: 1) effects on
pregnancy and birth outcomes, and 2) effects on female reproductive function. Evidence related to these
outcomes is described in Sections IS.7.3.7.1 and IS.7.3.7.3, respectively. Effects on development
(IS.7.3.7.2) and effects on male reproductive function (IS.7.3.7.3), for which evidence supports causal-
relationships with Pb exposure, are also discussed in more detail in the ensuing sections. Table IS-8A
through Table IS-8D provide a summary of the evidence from epidemiologic and animal toxicological
studies for reproductive and developmental effects, highlighting the recent evidence in comparison with
the evidence available in the 2013 Pb ISA.
IS.7.3.7.1 Effects on Pregnancy and Birth Outcomes
The 2013 Pb ISA concluded that the available evidence was "suggestive of a causal relationship
between Pb exposure and birth outcomes/' The causality determination was supported by associations
observed in epidemiologic studies of preterm birth and low birth weight/fetal growth. Notably, some
studies reported associations between Pb and low birth weight in studies that used postpartum maternal
bone Pb or air Pb concentrations. Although epidemiologic evidence was less consistent for associations
between low birth weight and maternal blood Pb measured during pregnancy or at delivery, or with Pb
measured in the umbilical cord and placenta, some inverse associations were observed between Pb
biomarker levels and birth weight or other measures of fetal growth. The effects of Pb exposure during
gestation in animal toxicological studies included similarly inconsistent findings, though most studies
reported reductions in birth weight of pups or litters when dams were treated with Pb. Thus, although the
evidence was inconsistent overall, there was some epidemiologic evidence supporting associations
between Pb exposure and preterm birth and low birth weight or fetal growth that was supported by
experimental animal evidence Pb-induced reductions in birthweight.
A recent quasi-experimental study used a difference-in-difference approach to demonstrate that
reducing potential exposures to airborne Pb reduced the risk of preterm birth and several other birth-
related outcomes. The study examined variation in potential airborne Pb exposure following the National
Association for Stock Car Auto Racing's (NASCAR's) deleading of racing fuel and reported that removal
of Pb from fuel was associated with increased birth weight as well as decreased probability of low birth
weight, preterm birth, and small for gestational age for children born to mothers living within 4,000m of a
racetrack relative to those residing at least 10,000m from the track. The difference-in-difference
methodology controls for time-varying confounders, removing biases from comparisons over time in the
treatment group that could be the result of trends due to other causes of the outcome. These findings
provide support for the effects of airborne Pb exposures on birth outcomes and are coherent with
experimental animal evidence from the 2013 Pb ISA indicating reductions in rodent birthweight
following Pb exposure. Additionally, there were a few high-quality epidemiologic studies that reported
associations with relevant BLLs and prenatal growth, birth defects, spontaneous abortion and pregnancy
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loss, and placental function, but the overall findings were inconsistent. There continue to be uncertainties
related to the specific biomarkers of exposure (maternal blood, maternal serum, maternal bone, maternal
erythrocytes, cord blood, cord blood serum, placental tissue) associated with pregnancy and birth
outcomes, the critical window of exposure, and potential confounding by co-occurring metals. Of note,
the cohorts in the recent epidemiologic literature would generally be expected to have had appreciable
past exposures to Pb; however, the extent to which adult BLLs in these cohorts reflect the higher exposure
histories is unknown as is the extent to which these past Pb exposures (magnitude, duration, frequency)
may or may not elicit effects on pregnancy and birth outcomes. Recent evidence from toxicological
studies mostly reported no effects of Pb across pregnancy and birth outcomes, but this may be due to the
exclusion of toxicological studies with exposures resulting in BLLs greater than 30 (ig/dL, indicating the
possibility that most pregnancy and birth outcomes are only affected in laboratory animals at levels higher
than most environmentally relevant Pb exposure levels.
In summary, recent epidemiologic studies expand on findings presented in the 2013 Pb ISA,
particularly supporting Pb effects on preterm birth and low birthweight and are coherent with previous
studies in experimental animals. The collective evidence is sufficient to conclude that there is likely to
be a causal relationship between Pb exposure and effects on pregnancy and birth outcomes.
Table IS-8A Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and reproductive and developmental
effects
Pregnancy and Birth Outcomes: Likely to Be Causal (IS.7.3.7.2 and Appendix 8.3)
Evidence from the 2013 Pb ISA
Epidemiologic evidence for pregnancy and birth outcomes
was inconsistent, but those that examined postpartum
maternal bone Pb or air Pb concentrations reported
associations with preterm birth and low birth weight.
Associations were less consistent for low birth weight with
maternal blood Pb levels (measured during pregnancy or at
delivery) or with Pb measured in the umbilical cord and
placenta. Most experimental animal studies reported
reductions in birth weight of pups or birth weight of litters
when dams were treated with Pb.
Evidence from the 2024 Pb ISA
A recent quasi-experimental study provides evidence
for decreased preterm birth rates following
NASCAR's phase out of leaded gasoline at races.
Recent evidence from experimental animal studies
generally reported no effects of Pb across pregnancy
and birth outcomes, but this may be due to the
exclusion of toxicological studies with exposures
resulting in BLLs greater than 30 |jg/dL.
BLL = blood lead level; mo = month(s); NASCAR = National Association for Stock Car Auto Racing; Pb = lead; yr = year(s).
IS.7.3.7.2 Effects on Development
The 2013 Pb ISA determined that the collective evidence was "sufficient to conclude that there is
a causal relationship between Pb exposures and developmental effects." This determination was based on
a strong body of evidence demonstrating delayed pubertal onset among males and females exposed to Pb.
Cross-sectional epidemiologic studies reported consistent associations between BLLs and delayed
pubertal onset (measured by age at menarche, pubic hair development, and breast development) among
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girls (ages 6-18 years) with mean and/or median concurrent BLLs of 1.2-9.5 (ig/dL. Although fewer
studies were conducted in boys, associations between BLLs and delayed puberty onset in boys (ages 8-
15 years) were observed in a longitudinal study and a few supporting cross-sectional studies (mean and/or
median BLLs of 3-9.5 (ig/dL). Limitations across most of the epidemiologic studies of BLLs and delayed
puberty included a lack of adjustment for nutritional factors as a potential confounder and the use of
cross-sectional study designs, which do not establish temporality. Additionally, because studies included
older children and adolescents who likely had higher earlier childhood Pb exposures, there is uncertainty
regarding the level, timing, frequency, and duration of Pb exposure that contributed to the observed
associations. Experimental animal studies demonstrate that puberty onset in both males and females is
delayed following exposure to Pb. Evidence for effects on postnatal growth was inconsistent.
Recent epidemiologic evidence continues to support an association between BLLs and delayed
pubertal onset in girls (Appendix 8.4.2) and boys (Appendix 8.4.3). Notably, recent studies observe more
consistent associations between Pb exposure and effects on puberty in girls. Although associations are
reported in populations with lower mean BLLs (0.65-6.57 (.ig/dL). uncertainty regarding the role of
potentially higher past exposures remains. Recent epidemiologic studies consider a wide range of
confounders, including height, weight, and body mass index (BMI), and some studies were conducted
among established longitudinal cohorts. No recent PECOS-relevant toxicological studies reported on the
effects of Pb on male or female puberty, though some studies provide evidence for the biological
plausibility of delayed pubertal onset. Specifically, Pb-induced disruptions of the hypothalamic-pituitary-
gonadal axis, steroidogenic enzymes, and their sex steroid products provide plausible pathways through
which Pb exposure could lead to the observed delays in pubertal onset reported in epidemiologic and
toxicological studies. Recent toxicological and epidemiologic evidence for effects on postnatal growth is
largely inconsistent, though epidemiologic studies that examined BLLs, as opposed to other biomarkers,
provide more consistent patterns of inverse associations between Pb exposure and height and weight in
children (8 months to 11 years).
Because of the strong body of evidence demonstrating delayed pubertal onset among males and
females exposed to Pb, the collective evidence is sufficient to conclude that there is a causal
relationship between Pb exposure and effects on development.
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Table IS-8B Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and reproductive and developmental
effects
Development: Causal Relationship (IS.7.3.7.2 and Appendix 8.4)
Evidence from the 2013 Pb ISA
Epidemiologic studies reported associations between
concurrent BLLs and delayed pubertal onset in boys and
girls. Associations were observed in children and
adolescents (6-18 yr) with low mean and/or median BLLs
(1.2-9.5 |jg/dL). A limitation across most of these studies is
their cross-sectional design, which does not establish
temporality between the exposure and outcome.
Additionally, there is uncertainty with regard to the exposure
frequency, timing, duration, and level that contributed to the
associations observed in these studies. Experimental animal
studies demonstrated that puberty onset in both males and
females is delayed with Pb exposure.
BLL = blood lead level; mo = month(s); Pb = lead; yr = year(s).
Evidence from the 2024 Pb ISA
Recent epidemiologic and toxicological evidence
continues to support Pb-related delays in pubertal
onset in boys and girls, including associations at
lower BLLs in the epidemiologic studies (0.65-
6.57 |jg/dL). Results from recent studies examining
the relationship between Pb exposure and postnatal
growth are inconsistent, though epidemiologic studies
that examined BLLs, as opposed to other biomarkers,
provide more consistent patterns of inverse
associations between Pb exposure and height and
weight in children (8 mo to 11 yr).
IS.7.3.7.3 Effects on Female Reproductive Function
The 2013 Pb ISA concluded that the available evidence was "suggestive of a causal relationship
between Pb exposure and female reproductive function." A small number of epidemiologic studies
reviewed in the 2013 Pb ISA reported associations with concurrent BLLs and altered hormone levels in
adults, but results were inconsistent, possibly due to the between study variation in hormones examined
and the timing of measurements as related to menstrual and lifecycles. There was additionally some
evidence of a potential inverse relationship between Pb exposure and female fertility, but findings were
again inconsistent. There were a number of study limitations in the epidemiologic evidence. The majority
of studies were cross-sectional and adjustment for potential confounders varied from study to study, with
some potentially important confounders, such as BMI, not included in all studies. Further, most of the
epidemiologic studies on female reproductive function reviewed in the 2013 Pb ISA had small sample
sizes and were generally conducted in women attending infertility clinics. Toxicological studies often
employed prenatal or early postnatal Pb exposures at relevant Pb levels and reported Pb-induced
decreases in ovarian antioxidant capacity, altered ovarian steroidogenesis, and aberrant gestational
hormone levels. Although epidemiologic and toxicological studies provide information on different
exposure periods, both types of studies, including some high-quality epidemiologic and toxicological
studies, supported the conclusion that Pb may affect some aspects of female reproductive function.
Recent studies expand on findings presented in the 2013 Pb ISA. The strongest line of evidence
comes from recent epidemiologic studies examining the relationship between Pb exposure and effects on
hormone levels and menstrual/estrous cyclicity (Table 8-1). Positive associations from a longitudinal
cohort between bone Pb, a biomarker of cumulative Pb exposure, and both earlier age at menopause and
risk of early menopause were supported by results from a cross-sectional NHANES study of concurrent
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exposure of blood Pb with earlier age at menopause. Additionally, recent epidemiologic studies found
consistent positive associations between blood Pb and follicle stimulating hormone and luteinizing
hormone in women who were postmenopausal. Although these studies are limited by their cross-sectional
study designs, they were conducted in well-established population-based surveys. These studies
considered a wider range of potential confounders compared to studies evaluated in the 2013 Pb ISA,
including coexposure to other metals, but not all studies adjusted for potentially important confounders
such as age at menarche, pregnancy history, oral contraceptive use, and female hormone use. As with
other studies in adults, the extent to which adult BLLs in these cohorts reflect potentially higher exposure
histories is undiscernible as is the extent to which these past Pb exposures (magnitude, duration,
frequency) may or may not elicit effects. While there were no recent PECOS-relevant toxicological
studies that examined the effects of Pb on hormone levels in females or menstrual or estrous cyclicity,
previous toxicological evidence supports epidemiologic study findings and indicate that Pb may disrupt
reproductive hormones and menstrual and estrous cyclicity in females. The collective evidence is
sufficient to conclude that there is likely to be a causal relationship between Pb exposure and female
reproductive function.
Table IS-8C Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and reproductive and developmental
effects
Female Reproductive Function: Likely to Be Causal (IS.7.3.7.2 and Appendix 8.5)
Evidence from the 2013 Pb ISA
A limited number of experimental animal studies conducted
in nonhuman primates and rodents reported disrupted
menstrual or estrous cyclicity and reduced progesterone
following high levels of exposure to Pb (BLLs: 44-
264 |jg/dL), although another nonhuman primate study with
lower BLLs than the other studies (<40 |jg/dL) reported no
effects on menstrual cyclicity. Cross-sectional epidemiologic
studies with inconsistent adjustment for important
confounders reported some associations between
concurrent BLLs and altered hormone levels in women
attending infertility clinics.
BLL = blood lead level; mo = month(s); Pb = lead; yr = year(s).
Evidence from the 2024 Pb ISA
A recent cohort study reported associations between
bone Pb, a biomarker of cumulative Pb exposure, and
both earlier age at menopause and risk of early
menopause. Cross-sectional studies provided
supporting evidence of positive associations between
concurrent exposure of blood Pb with earlier age at
menopause. While there were no recent
PECOS-relevant experimental animal studies that
examined the effects of Pb on menstrual or estrous
cyclicity, the results from recent epidemiologic studies
are coherent with previous animal toxicological
evidence.
IS.7.3.7.4 Effects on Male Reproductive Function
In the 2013 Pb ISA, the evidence was "sufficient to conclude that there is a causal relationship
between Pb exposures and male reproductive function." Key evidence was provided by toxicological
studies in rodents, nonhuman primates, and rabbits showing detrimental effects on semen quality, sperm,
and fecundity/fertility with supporting evidence in epidemiologic studies of associations between BLLs
and detrimental effects on sperm. Animal exposures resulting in BLLs from 5-43 (ig/dL induced lower
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sperm quality and sperm production rate, sperm DNA damage, and histological or ultrastructural damage
to the male reproductive organs. These effects were found in animals exposed to Pb for 1 week to
3 months during peripuberty or as adults. Pb exposure of male rats also resulted in subfecundity in female
mates and lower fertilization of eggs in vitro. Detrimental effects of Pb on sperm were observed in
epidemiologic studies with concurrent BLLs of 25 (ig/dL and greater among occupationally exposed men;
however, these studies were limited because of their lack of consideration of potential confounding
factors, including occupational exposures other than Pb. A smaller number of epidemiologic studies
among men with lower Pb biomarker levels were limited to fertility clinic studies that may lack
generalizability. Additionally, because of uncertainty regarding greater exposure to Pb earlier in life in
these populations, the extent to which adult BLLs in these cohorts reflect potentially higher exposure
histories as well as the extent to which these past Pb exposures (magnitude, duration, frequency) may or
may not elicit effects is not discernible from the epidemiologic evidence. Biological plausibility for the
observed associations was provided by animal toxicological studies that demonstrated Pb-induced
oxidative stress within the male sex organs, increase apoptosis of spermatocytes and germ cells, and
impaired germ cell structure and function.
Recent epidemiologic evidence continues to support an association between BLLs and decreased
sperm/semen production, quality, and function. Results from analyses using other Pb biomarkers,
including plasma, semen, and seminal fluid, were inconsistent. The evaluated studies were cross-sectional
and conducted in males attending fertility clinics, which may have resulted in selection bias and limits the
generalizability of the results. The studies were also limited by concurrent measurement of exposure and
outcome, examination of different seminal parameters, and small sample sizes. Despite these limitations,
a wide variety of potential confounders were considered, including adjustment for hormone levels, which
could potentially impact sperm/semen production, quality, and function. Recent toxicological studies
generally report that Pb exposure alters some aspects of sperm or semen quality, such as sperm density,
motility, morphology, and viability, especially studies that include dosing during developmental periods
or for periods 30 days or longer. The strongest line of evidence, including potential biologically plausible
pathways, were reported for effects on sperm/semen production, quality, and function, while evidence for
other effects on male reproductive function, including hormone levels, male fertility, and morphology and
histology of male sex organs is either limited in quantity and/or inconsistent. Overall, the collective
evidence is sufficient to conclude that there is a causal relationship between Pb exposure and male
reproductive function.
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Table IS-8D Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and reproductive and developmental
effects
Male Reproductive Function: Causal Relationship (IS.7.3.7.3 and Appendix 8.6)
Evidence from the 2013 Pb ISA
Animal toxicological studies in rodents, nonhuman primates,
and rabbits reported that Pb exposures resulting in BLLs
from 5-43 |jg/dL induced lower sperm quality and sperm
production rate, sperm DNA damage, and histological or
ultrastructural damage to the male reproductive organs.
These effects were found in animals exposed to Pb for 1 wk
to 3 mo during peripuberty or as adults. There was some
supporting epidemiologic evidence, but most studies
examined occupational^ exposed men with high BLLs
(>25 |jg/dL) and included limited control for potential
confounders.
Evidence from the 2024 Pb ISA
Recent epidemiologic studies reported consistent
associations between BLLs and decreased
sperm/semen production and quality. Results were
inconsistent in studies that measured Pb in seminal
fluid or seminal plasma. Epidemiologic studies also
provided initial evidence of an association between
BLLs and increased testosterone and morphological
changes in male sex organs. The epidemiologic
studies evaluated include non-occupationally
exposed men with lower Pb exposures than studies
included in the 2013 Pb ISA. Recent toxicological
evidence is consistent with findings from the 2013 Pb
ISA.
BLL = blood lead level; ISA = Integrated Science Assessment; mo = month(s); Pb = lead; wk = week(s); yr = year(s).
IS.7.3.8 Musculoskeletal Effects
The 2013 Pb ISA concluded that "a causal relationship is likely to exist between Pb exposure and
effects on bone and teeth." In order to be more inclusive of other health effects related to bone and teeth
(e.g., muscles, joints, and cartilage), this ISA expands the considered health outcomes to include effects
on the entire musculoskeletal system. A summary of the evidence available in the 2013 Pb ISA as well as
evidence from recent studies is provided in Table IS-9. Recent epidemiologic evidence continues to
support an association between Pb exposure and effects on bone (e.g., increased prevalence of
osteoporosis) and teeth (i.e., increased prevalence and incidence of dental caries and tooth loss in children
and adults). There is also an emerging area of research on osteoarthritis, an endpoint that was not
discussed in the 2013 Pb ISA. A few recent cross-sectional studies reported positive associations between
BLLs and symptomatic and radiographic osteoarthritis and some biomarkers of joint tissue metabolism.
The epidemiologic evidence base includes a larger number of studies and adult populations with lower
mean, median, or geometric mean BLLs than studies included in the 2013 Pb ISA (1.03 to 4.44 (ig/dL).
Despite the evidence for associations at relatively low BLLs, these musculoskeletal outcomes were most
often examined in adults who have been exposed to higher levels of Pb earlier in life, the extent to which
adult BLLs in these cohorts reflect potentially higher exposure histories as well as the extent to which
these past Pb exposures (magnitude, duration, frequency) may or may not elicit effects is not discernible
from the epidemiologic evidence. Additionally, although the recent epidemiologic evidence is consistent
with the findings highlighted in the 2013 Pb ISA, recent studies do not thoroughly address the unclear
temporality of exposure and outcome resulting from mostly cross-sectional study designs. This
uncertainty is particularly important for studies examining benchmark dose (BMD) and osteoporosis due
to the possibility of reverse causality, where the observed associations could be driven by higher BLLs
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due to increased bone turnover in individuals with low BMD or osteoporosis. The toxicologic data
support Pb-induced alterations in multiple aspects of bone, teeth, and joint maintenance. For skeletal
bones, shift in the balance between bone building osteoblasts and bone resorbing osteoclasts could be
responsible for delayed bone growth and increased bone degeneration seen in epidemiologic studies. In
teeth and joints, Pb appears to suppress the synthesis of cellular matrix proteins important for joint
maintenance and enamel formation which could plausibly contribute to the osteoarthritic and dental
effects seen in some epidemiologic studies. Overall, the collective evidence is sufficient to conclude
that there is likely to be a causal relationship between Pb exposure and musculoskeletal effects.
Table IS-9 Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and musculoskeletal effects
Musculoskeletal Effects: Likely to Be Causal (IS.7.3.8 and Appendix 9.5)
Evidence from the 2013 Pb ISA
Strong toxicological evidence evaluated in the 2013 Pb ISA
and the 2006 Pb AQCD (U.S. EPA. 2006) demonstrates
effects in bone and teeth in animals following Pb exposure.
Exposure of animals to Pb during gestation and the
immediate postnatal period was reported to significantly
depress early bone growth with concentration-dependent
trends. Systemic effects of Pb exposure included disruption
of bone mineralization during growth, alterations in bone cell
differentiation and function due to alterations in plasma
levels of growth hormones and calcitropic hormones such as
1,25-dihydroxyvitamin D3, effects on Ca2+- binding proteins,
and increases in Ca2+ and phosphorus concentrations in the
bloodstream. As in bone, Pb was found to easily substitute
for Ca2+ in the teeth following exposure and was taken up
and incorporated into developing teeth in experimental
animals. These findings were coherent with results from a
small body of epidemiologic studies that provided consistent
evidence of associations between Pb biomarker levels and
various effects on bone and teeth after adjusting for potential
confounding by age and SES-related factors.
Evidence from the 2024 Pb ISA
Recent epidemiologic studies continue to support
associations between Pb exposure and effects on
bone in adults and teeth in children and adults. The
recent epidemiologic evidence is mostly from cross-
sectional studies and does not thoroughly address
the temporality of exposure and outcome.
Additionally, uncertainty remains concerning the Pb
exposure level, timing, frequency, and duration
contributing to the observed associations in adult
populations. Recent toxicological evidence is limited,
but consistent with findings from the 2013 Pb ISA and
coherent with the epidemiologic evidence.
AQCD = Air Quality Criteria Document; Ca2+ = calcium ion; ISA = Integrated Science Assessment; Pb = lead;
SES = socioeconomic status.
IS.7.3.9 Mortality
In the 2013 Pb ISA (U.S. EPA. 2013a). the strongest evidence for Pb-associated mortality was
from studies examining cardiovascular mortality. The evidence did not provide strong support for Pb-
associated mortality other than through cardiovascular pathways, and very few studies examined total
(nonaccidental) mortality. For these reasons, the 2013 Pb ISA evaluated studies of all-cause mortality
together with studies examining cardiovascular mortality, and these studies were all included within the
CVD chapter. Although this evidence contributed to the "causal relationship" between Pb exposure and
CHD, there was no distinct causality determination for total or cause-specific mortality. A small number
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of studies evaluated in the 2013 Pb ISA reported consistently positive associations between Pb
biomarkers and total mortality. This evidence was further supported by consistent evidence of positive
associations between BLLs and cardiovascular mortality in NHANES cohorts, including some studies
that controlled for a wide range of potential confounders, tested for interactions between confounders and
BLL, included evaluations of C-R relationships and extensive analysis of model evaluations, and
examined specific causes of CVD mortality. In addition, an analysis of the NAS reported an association
between bone Pb, a metric of cumulative Pb exposure, and increased total and cardiovascular mortality in
older male veterans.
Several recent epidemiologic studies build upon evidence from the 2013 Pb ISA and provide
largely consistent evidence of an association between biomarkers of Pb exposure and total and
cardiovascular mortality (Table IS-10). A recent quasi-experimental study comparing time periods prior
to and after the phaseout of leaded gasoline in professional racing series (i.e., NASCAR and the
Automobile Racing Club of America [ARCA]) observed a decline in mortality rates in race counties
relative to control counties following the phaseout of leaded gasoline. The novel study design utilized in
this analysis is able to reduce concerns of potential confounding under a set of well-reasoned, but
untestable assumptions. Other recent studies include nationally representative adult populations with low
BLLs (mean <2.5 |ig/dL). including an extended analysis of the NHANES III cohort. Notably, these
analyses include study populations that were born prior to the phaseout of leaded gasoline and therefore
likely had much higher past Pb exposures. Thus, the extent to which adult BLLs in these cohorts reflect
potentially higher exposure histories as well as the extent to which these past Pb exposures (magnitude,
duration, frequency) may or may not elicit effects is not discernible from the epidemiologic evidence.
Studies that examined multiple causes of mortality in the same cohort generally reported effect estimates
that were notably smaller in magnitude for total mortality compared to cardiovascular mortality. This
suggests that the total mortality results may in large part be driven by the association between BLLs and
cardiovascular mortality. There is extensive epidemiologic and toxicological evidence indicating
pathways by which exposure to Pb could plausibly progress from initial events to endpoints relevant to
the cardiovascular system, such as hypertension, exacerbation of IHD, and potential MI or stroke.
Because cardiovascular morbidity, which comprises 33% of total (nonaccidental) mortality, is the most
common contributor to total mortality (NHLBI. 2017). the progression demonstrated in the available
evidence for cardiovascular morbidity supports potential biological pathways by which Pb exposure could
result in cardiovascular mortality. There is also very limited evidence that Pb exposure is positively
associated with other causes of mortality, including Alzheimer's disease (AD) and infection. Biological
plausibility for these outcomes is demonstrated by pathways leading from Pb exposure to
neurodegenerative disease (Appendix 3.3) and immunosuppression (Section IS.7.3.5), respectively.
However, although there is toxicological evidence that developmental exposure to Pb increases the
expression of proteins related to AD, the epidemiologic evidence relating Pb exposure to incident AD
remains limited. A few uncertainties remain in the evidence base, including a limited number of
independent studies (i.e., from non-overlapping study populations), and uncertainty regarding to the
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specific timing, duration, frequency, and level of Pb exposure that contributed to the observed
associations.
Given the strong epidemiologic evidence for Pb-associated all-cause and cardiovascular mortality
and strong supporting evidence for Pb-associated cardiovascular effects, there is sufficient evidence to
conclude that there is a causal relationship between Pb exposure and total (nonaccidental)
mortality.
Table IS-10 Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and total (nonaccidental) mortality
Total (Nonaccidental) Mortality: Causal (IS.7.3.9 and Appendix 9.8)
Evidence from the 2013 Pb ISA
Consistent evidence of positive associations between BLLs
and total and cardiovascular mortality observed in NHANES
cohorts, including some studies that controlled for a wide
range of potential confounders. In addition, an analysis of
the NAS reported an association between bone Pb, a metric
of cumulative Pb exposure, and increased total and
cardiovascular mortality in older male veterans.
Evidence from the 2024 Pb ISA
Recent epidemiologic studies build upon evidence
from the 2013 Pb ISA and provide largely consistent
evidence of an association between biomarkers of Pb
exposure and total and cardiovascular mortality.
Recent studies include a quasi-experimental study
and nationally representative populations with low
BLLs (mean <2.5 |jg/dL). Uncertainties remain
regarding the specific timing, duration, frequency, and
level of Pb exposure that contributed to the observed
associations.
BLL = blood lead level; ISA = Integrated Science Assessment; NAS = Normative Aging Study; NHANES = National Health and
Nutrition Examination Survey; mo = month(s); Pb = lead; yr = year(s).
IS.7.3.10 Cancer
The 2013 Pb ISA concluded that "a causal relationship is likely to exist between Pb exposure and
cancer." This determination was based on strong evidence from animal toxicological studies
demonstrating effects of Pb on cancer, genotoxicity, or epigenetic modification (Table IS-11).
Carcinogenicity in animal toxicological studies with relevant routes of Pb exposure were reported in the
kidneys, testes, brain, adrenals, prostate, pituitary, and mammary gland, albeit at high doses of Pb.
Epidemiologic studies of cancer incidence and mortality reported inconsistent results; one strong
epidemiologic study demonstrated an association between BLLs and increased cancer mortality, but other
studies reported weak (i.e., small magnitude and/or imprecise 95% CIs) or null associations. The
consistent evidence indicating Pb-induced carcinogenicity in animal models was substantiated by the
mode of action findings from multiple high-quality toxicological studies in animal and in vitro models
from different laboratories.
There are no recent toxicological studies conducted at concentrations deemed relevant to this ISA
(i.e., BLLs <30 (ig/dL). Recent in vitro studies add to our understanding of how Pb exposures may
activate the mechanistic pathways that can result in cancer, including evidence for Pb activation of
mechanistic pathways mediated by oxidative stress, genotoxicity, and inflammation, as well as changes in
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cell cycle regulatory genes, epigenetics, apoptosis, and necrosis (Appendix 10.3). Additionally, new areas
of research involving matrix metalloproteinases and metallothioneins have emerged and provide evidence
of other potential mechanistic pathways through which Pb exposure could contribute to cancer. In the
absence of any new cancer bioassay studies using animal models, uncertainty remains regarding the
carcinogenic potential of low levels of Pb exposure. Recent epidemiologic evidence does little to address
this uncertainty. Similar to the epidemiologic evidence evaluated in the 2013 Pb ISA, recent
epidemiologic studies observed inconsistent associations between Pb exposure and overall cancer
mortality (Appendix 10.4.2). A limited number of recent studies evaluating Pb exposure and site-specific
cancers is also inconsistent. The small body of evidence across various site-specific cancer endpoints
limits the ability to judge coherence and consistency across these studies. In general, recent studies
control for a wide range of potential confounders, but studies were limited by a small number of cases
resulting in limited power to detect an association, a relatively short time period between exposure and
outcome, potential differences in Pb exposure histories based on study location, and the use of different
biomarkers of exposure. Additionally, when associations were observed, study populations most often
included adults who have been exposed to higher levels of Pb earlier in life, which produces uncertainty
regarding the Pb exposure level, timing, frequency, and duration contributing to the observed
associations.
Given the strong support from cancer bioassay studies using animal models with high exposure
concentrations and in vitro studies of mechanistic pathways indicating the carcinogenic potential of Pb
exposures, the collective evidence is sufficient to conclude that there is likely to be a causal
relationship between Pb exposure and cancer.
Table IS-11 Summary of evidence from epidemiologic and animal toxicological
studies on Pb exposure and cancer
Cancer: Likely to Be Causal (IS.7.3.10 and Appendix 10)
Evidence from the 2013 Pb ISA
Toxicological studies consistently reported cancer incidence
following chronic exposure (i.e., 18 mo or 2 yr) to high
concentrations of Pb, such as 10,000 ppm Pb acetate in diet
or 2,600 ppm Pb acetate in drinking water. High-quality
toxicological studies in animal and in vitro models from
different laboratories also provided a biologically plausible
pathway through which Pb exposure could lead to cancer.
Epidemiologic studies of cancer incidence and mortality
reported inconsistent results.
Evidence from the 2024 Pb ISA
No recent cancer bioassay studies using animal
models with relevant exposure levels are available. In
vitro studies provide additional evidence supporting
the Pb-induced activation of diverse mechanistic
pathways that are typically associated with
carcinogenesis. Recent epidemiologic studies add to
the inconsistent epidemiologic evidence of an
association between Pb exposure and cancer
mortality.
ISA = Integrated Science Assessment; mo = month(s); Pb = lead; yr = year(s).
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IS.7.4
At-Risk Populations
Interindividual variation in exposure or human responses to ambient air pollution can result in
some groups or lifestages being at increased risk for health effects. The NAAQS are intended to protect
public health with an adequate margin of safety. In so doing, protection is provided for both the
population as a whole and those at increased risk for health effects in response to exposure to a criteria air
pollutant [e.g., Pb; see Preamble (U.S. EPA. 2015)1. There is interindividual variation in both
physiological responses and exposure to Pb in the environment. The scientific literature has used a variety
of terms to identify factors and subsequently populations or lifestages that may be at increased risk of an
air pollutant-related health effect, including susceptible, vulnerable, sensitive, at-risk, and response-
modifying factors (U.S. EPA. 2015). Acknowledging the inconsistency in definitions for these terms
across the scientific literature and the lack of a consensus on terminology in the scientific community, "at-
risk" is the all-encompassing term used in ISAs for groups with specific factors that increase the risk of an
air pollutant (e.g., Pb)-related health effect in a population, as initially detailed in the 2013 Pb ISA (U.S.
EPA. 2013a). Therefore, this ISA takes an inclusive and all-encompassing approach and focuses on
identifying those populations or lifestages potentially "at risk" of a Pb-related health effect.
As discussed in the Preamble (U.S. EPA. 2015). the risk of health effects from exposure to Pb
may be modified as a result of intrinsic (e.g., preexisting disease, genetic factors) or extrinsic factors
(e.g., sociodemographic or behavioral factors), differences in internal dose, or differences in exposure to
Pb in the environment. Some factors may lead to a reduction in risk and are recognized as such during the
evaluation. However, to inform decisions on the NAAQS, this ISA focuses on identifying those
populations or lifestages at greater risk. While a combination of factors (e.g., residential location and
SES) may increase the risk of Pb-related health effects in portions of the population, information on the
interaction among factors remains limited. Thus, this ISA characterizes the individual factors that
potentially result in increased risk for Pb-related health effects [see Preamble (U.S. EPA. 2015)1.
IS.7.4.1 Approach to Evaluating and Characterizing the Evidence for At-Risk Factors
The ISA identifies and evaluates factors that may increase the risk of a population or specific
lifestage to a Pb-related health effect; this approach is described in detail in the Preamble (U.S. EPA.
2015) and is illustrated in Table IS-12. Whereas Appendices 3-10 include a discussion of some
populations and lifestages in order to explicitly characterize the causal nature between Pb biomarkers of
exposure and health effects based on the body of evidence (e.g., children, minority populations), this
section focuses on summarizing evidence that can inform the identification of such populations and
lifestages.
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Table IS-12 Characterization of evidence for factors potentially increasing the
risk for Pb-related health effects
Classification Health Effects
Adequate evidence There is substantial, consistent evidence within a discipline to conclude that a factor results
in a population or lifestage being at increased risk of air pollutant-related health effect(s)
relative to some reference population or lifestage. Where applicable, this evidence includes
coherence across disciplines. Evidence includes multiple high-quality studies.
Suggestive evidence The collective evidence suggests that a factor results in a population or lifestage being at
increased risk of air pollutant-related health effect(s) relative to some reference population
or lifestage, but the evidence is limited due to some inconsistency within a discipline or,
where applicable, a lack of coherence across disciplines.
Inadequate evidence The collective evidence is inadequate to determine whether a factor results in a population
or lifestage being at increased risk of air pollutant-related health effect(s) relative to some
reference population or lifestage. The available studies are of insufficient quantity, quality,
consistency, and/or statistical power to permit a conclusion to be drawn.
Evidence of no effect There is substantial, consistent evidence within a discipline to conclude that a factor does
not result in a population or lifestage being at increased risk of air pollutant-related health
effect(s) relative to some reference population or lifestage. Where applicable, the evidence
includes coherence across disciplines. Evidence includes multiple high-quality studies.
The evidence evaluated in this section includes relevant studies discussed in Appendix 3-
Appendix 10 of this ISA and builds on the evidence presented in the 2013 Pb ISA (U.S. EPA. 2013a).
Using the approach developed in previous ISAs, (U.S. EPA. 2020. 2016a. 2013a. b) recent evidence is
integrated across scientific disciplines and health effects, and where available, with information on
exposure and dosimetry. In evaluating factors and population groups, greater emphasis is placed on the
evidence for those health outcomes for which a "causal" or "likely to be causal" relationship is concluded
in Appendix 3-Appendix 10 of this ISA (see Section IS.7.3).
As discussed in the Preamble (U.S. EPA. 2015). consideration of at-risk populations includes
evidence from epidemiologic and animal toxicological studies, in addition to relevant exposure-related
information. Regarding epidemiologic studies, the evaluation focuses on those studies that include
stratified analyses to compare populations or lifestages exposed to similar air pollutant concentrations
within the same study design along with consideration of the strengths and limitations of each study.
Other epidemiologic studies that do not stratify results but instead examine a specific population or
lifestage can provide supporting evidence for the pattern of associations observed in studies that formally
examine effect measure modification.
Effect modification occurs when the effect of interest differs between subgroups or strata
(Rothman et al.. 2012). When a risk factor is an effect modifier, it changes the magnitude of the
association between exposure to Pb and the outcome of interest across those strata or subgroups. For
example, the presence of a preexisting disease or indicator of low socioeconomic status (SES)
(e.g., educational attainment, household income) may act as an effect modifier if it is associated with
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increased or decreased risk of Pb-related health effects. Thus, evidence of effect modification can help
identify at-risk factors or potentially at-risk populations.5
Inference can be particularly strong from studies that consider the potential impacts of effect
modification, especially when the modifying factors are coherent with information from other lines of
evidence regarding the biological pathways connecting Pb exposures with particular health effects.
Traditional modeling approaches, such as stratification and interaction terms, can identify individual
effect modifiers (e.g., age groups or preexisting diseases), while emerging modeling approaches can
identify a set of complex moderation functions. Preference in this section is given to studies articulating
and justifying assumptions of effect modification and to studies with appropriate diagnostics
(e.g., multiple comparisons) accounting for potentially spurious findings.
Similar to the characterization of evidence in Appendix 3-Appendix 10, the greatest emphasis is
placed on patterns or trends in results across studies. Experimental studies in animals that focus on
factors, such as genetic background or preexisting disease, are evaluated because they provide coherence
and can support the biological plausibility of effects observed in epidemiologic studies. Also evaluated
are studies examining whether factors may result in differential exposure to Pb and subsequent increased
risk of Pb-related health effects. Additionally, physiologic factors that may influence the internal
distribution of Pb are also considered. Conclusions are made with respect to whether a specific factor
increases the risk of a Pb-related health effect based on the characterization of evidence using the
framework detailed in Table III of the Preamble (U.S. EPA, 2015), and presented in Table IS-12.
IS.7.4.2 Summary of Population Characteristics and Other Factors Potentially Related
to Increased Risk of Pb-Related Health Effects
The 2013 Pb ISA (U.S. EPA. 2013a) concluded that there was adequate evidence to classify
children, minority populations, individuals in proximity to Pb sources, individuals living in residences
with factors contributing to increased house dust Pb levels, and those with a certain nutritional status as
populations at increased risk of Pb-related health effects. These conclusions were based on the
consistency in findings across studies, as well as on coherence of results from different scientific
disciplines. Some populations may be at increased risk of Pb-related health effects mostly due to
increased Pb exposure. Recent studies provide additional evidence that minority populations, children,
those in proximity to Pb sources, and those with certain nutritional excesses or deficiencies are at
5Effect modification may also inform causality determinations in several ways. Consistent evidence that at least one
population subgroup is at risk of a Pb-related health effect provides strong support for causality determinations.
Evidence for effect modification can also explain heterogeneity in results across studies, which could reduce
uncertainties regarding inconsistent evidence. Finally, effect modification can provide supporting information on
mechanisms (e.g., genetic polymorphisms or microbiome profiles) contributing to Pb-related health effects. Notably,
the lack of evidence for effect modification where there is otherwise evidence of a Pb-related health effect in the
general population does not weaken the overall evidence supporting a causality determination.
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increased risk for Pb-related health effects. There is relatively little recent evidence to add to the evidence
presented in the 2013 Pb ISA regarding individuals living in areas with certain residential factors
(Table IS-13).
Several recent large epidemiologic studies, including some longitudinal studies, evaluated health
effects among certain racial/ethnic groups or stratified results by race/ethnicity. Results from these studies
expand the current knowledge base from the 2013 Pb ISA to provide further support of the relationship
between Pb biomarkers and health effects (mainly increased concurrent BP and hypertension) among
Black and Asian populations. However, there remains uncertainty regarding the level, timing, frequency,
and duration of Pb exposure contributing to the observed associations. Similarly, recently available
evidence among children further elucidates the increased risk children can experience from elevated
exposures to Pb. Additionally, those living in proximity to a Pb sources (e.g., industrial sources of Pb) are
not only at increased risk of elevated Pb biomarker levels, due to increased Pb exposure, but also
increased risk of negative Pb-related health outcomes, as was demonstrated in the 2013 Pb ISA. Lastly,
the recent evidence further supports and adds to the collective evidence presented in the 2013 Pb ISA that
the presence of absence of certain nutrients (e.g., reduced intake of Ca2+ and Fe) may increase Pb-related
health effects, while other nutrient deficiencies or surpluses may decrease the risk of a Pb-related health
effect among certain populations.
Since the 2013 Pb ISA, recent research has expanded the evidence bases for several factors,
which were originally classified as providing suggestive evidence of a population or lifestage that
increases the risk of Pb-related health effects. Specifically, at the time of the 2013 Pb ISA there were a
limited number of studies that evaluated genetic variants in relation to the effects of Pb exposure on a
population. However, recent studies consider several additional genetic variants, and evidence collected
as a whole further elucidates differential effects among certain segments of the population with genetic
variants. Additionally, more evidence is available related to the impacts of stress on the health effects of
Pb exposure. Taken together, recent studies, in combination with studies evaluated in the 2013 Pb ISA,
provide adequate evidence that high stress levels modify the associations between Pb exposure and health
effects.
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Table IS-13 Summary of evidence for population characteristics and other
factors potentially related to increased risk of Pb-related health
effects
Conclusions from 2013 Pb ISA
Conclusions from the 2024 Pb ISA
Adequate evidence (2024 Pb ISA)
Race/ethnicity Compared with white populations,
minority populations were observed to
be more at risk of Pb-related health
effects. Studies of race/ethnicity
provide adequate evidence that
race/ethnicity is an at-risk factor based
on the higher exposure observed
among minority populations and some
modification observed in studies of
associations between Pb levels and
health effects.
Recent exposure studies demonstrate that non-
Hispanic Black children consistently have higher
than average BLLs, particularly when compared
with Hispanic and non-Hispanic white children,
even though overall BLLs are dropping. Recent,
large epidemiologic studies conducted in the
United States expand upon previous evidence
indicating that race/ethnicity is an effect measure
modifier for Pb-related health outcomes.
Childhood
In consideration of the evidence base
(e.g., stratified and longitudinal
analyses) and integrating across
disciplines of toxicokinetics, exposure,
and health, there is adequate evidence
to conclude that children are an at-risk
population.
Recent evidence supports previous conclusions
and extends findings among different childhood
age groups.
Proximity to Pb Epidemiologic studies report consistent
sources positive associations between
increased Pb exposure and associated
health effects among those in proximity
to Pb sources, including areas with
large industrial sources.
Recent epidemiologic evidence further supports
prior conclusions for both increased exposure and
increased risk of health effects in proximity to Pb
sources.
Nutrition
Epidemiologic and toxicologic studies
provide consistent evidence that
certain nutritional factors can increase
or decrease the association between
Pb exposure and certain Pb-related
health effects.
Epidemiologic and toxicologic studies provide
consistent evidence that certain nutritional factors
can increase or decrease the association
between Pb exposure and certain Pb-related
health effects.
Residential factors
Findings suggest positive associations
between increased blood Pb and
increased house dust Pb levels.
Recent information does not inform or change
prior conclusions.
Genetics
Few genetic variants have been
observed in epidemiologic and
controlled human exposure studies to
affect the risk of Pb-related health
outcomes and support is provided by
animal toxicological studies of genetic
factors.
Additional genetic variants, epigenetic
modifications, and gene expression factors have
been found to interact with Pb-related health
outcomes.
Stress
Stress was evaluated as a factor that
potentially increases the risk of Pb-
related health effects (e.g., cognitive
function in adults and hypertension),
and while limited by the small number
of epidemiologic studies, increased
stress was observed to exacerbate the
effects of Pb. Toxicological studies
supported this finding.
Recent evidence informs prior conclusions and
extends the results to children. Studies observed
that high levels of maternal stress exacerbated
the effect of prenatal Pb exposure on several
neurodevelopmental domains, including
language. Toxicological studies provide support
for the interaction between maternal stress and
Pb-related cognitive effects by sex.
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Conclusions from 2013 Pb ISA
Conclusions from the 2024 Pb ISA
Suggestive evidence (2024 Pb ISA)
Older adulthood Evidence, based on limited
Recent information does not inform or change
epidemiologic evidence but support
prior conclusions.
from toxicological studies and
differential exposure studies, is
suggestive that older adults are
potentially at risk of Pb effects.
However, there are uncertainties
related to the exposure profile
associated with the effects in older
populations.
Sex Potential evidence suggests that
Recent evidence informs prior conclusions, but
adolescent and adult males typically
still contains inconsistencies in presented results.
demonstrate higher BLLs, although
evidence regarding health outcomes is
limited due to inconsistencies in
whether males or females are at
greater risk of certain outcomes in
relation to Pb
Preexisting disease There are a limited number of
Recent information does not inform or change
epidemiologic studies that suggest
prior conclusions.
preexisting diabetes modifies Pb
effects on specific health effects
(e.g., renal function or cardiovascular
outcomes)
SES Studies of SES and its relationship with
Recent information does not inform or change
Pb-related health effects are few and
prior conclusions.
report inconsistent findings regarding
low SES as a potential at-risk factor.
Overall, the evidence is suggestive that
low SES is a potential at-risk factor for
Pb-related health effects.
Other metals High levels of other metals, such as Cd
Limited recent evidence informs prior conclusions.
and Mn, were observed to result in
Hg and As were also found to interact with Pb-
greater effects for the associations
related cognitive functions.
between Pb and various health
endpoints (e.g., renal function,
cognitive function in children), but
overall, the evidence was limited.
Inadequate evidence (2024 Pb ISA)
Smoking status There are a limited number of studies
Recent information does not inform or change
and insufficient coherence for
prior conclusions.
differences in Pb-related health effects
by smoking status.
BMI A small number of studies provide
Recent evidence suggests modification of Pb-
inadequate evidence that there may be
related health effects by overweight status.
BMI-related increase in risk of Pb-
related health effects for some
outcomes.
Alcohol consumption A small number of studies provide
Recent information does not inform or change
inadequate evidence that there may be
prior conclusions.
alcohol-related increases in Pb-related
health effects for some outcomes.
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Conclusions from 2013 Pb ISA
Conclusions from the 2024 Pb ISA
Maternal self-esteem A small number of studies related to Recent information does not inform or change
the relationship between Pb exposure prior conclusions,
and infant development suggested that
maternal self-esteem modified the
association, but the results were
inconsistent, especially across other
health outcomes.
Cognitive reserve Limited epidemiologic evidence Recent information does not inform or change
suggests that cognitive reserve may prior conclusions,
differentially impact the association
between Pb exposure and Pb-related
health outcomes. No additional
evidence from the 2013 Pb ISA
expanded the assessment of this
factor.
As = arsenic; BLL = blood lead level; BMI = body mass index; Cd = cadmium; Mn = manganese; Hg = mercury; ISA = Integrated
Science Assessment; Mn = manganese; Pb = lead; SES = socioeconomic status.
IS.7.4.2.1 Race/Ethnicity
Race is widely acknowledged to be a social construct, not a fixed biological (Pavnc-Sturgcs et al..
2021). Observed differences in exposures and/or outcomes across racial groups, therefore, are likely to
reflect race as a proxy measure for a complex set of factors that result from these societal constructs
(e.g., nutrition, housing opportunity, access/barriers to health care). This ISA evaluates and synthesizes
existing research on the health and welfare effects of exposure to Pb, and many studies evaluated herein
examine racial disparities in environmental exposure and human health, but do not empirically assess the
underlying complexities that contribute to said disparities. Identifying racial disparities is an important
step in recognizing populations at increased risk to the health effects of Pb but should also be considered
in the context of the specific underlying factors that might explain these differences in exposures and/or
outcomes. This section describes racial and ethnic disparities in Pb exposure and health effects, while
some of the ensuing sections address other factors that may be impacted by social constructs of race
(i.e., proximity to sources, nutrition, and stress).
Historically, racial and ethnic differences in exposures to environmental Pb have been evident.
Both the 2006 Pb AQCD and the 2013 Pb ISA presented consistent evidence that Black populations have
historically had relatively higher blood and bone Pb levels compared with white and other minority
populations. While the 2013 Pb ISA reported that racial and ethnic gaps in mean blood and bone Pb levels
have gradually narrowed over time, Black populations continue to typically have higher Pb exposures and
body burdens compared with white populations. Recent evidence from 2011-2018 NHANES cycles
indicates that non-Hispanic Black populations generally had BLLs higher than the national average, but in
more recent years, average BLLs in non-Hispanic Black populations were lower than in non-Hispanic
white populations (Appendix 2.4). Moreover, in some years, Asian populations had the highest mean
BLLs when compared with other racial/ethnic groups. Nonetheless, non-Hispanic Black children are
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consistently the group with the highest BLLs, although both overall differences and differences among
groups are declining.
The 2013 Pb ISA concluded that minority populations, specifically non-Hispanic Black
populations, are at an increased risk of health effects related to Pb exposure, compared with white
populations. This conclusion is supported by several longitudinal and cross-sectional analyses. Recent
large epidemiologic studies conducted in the United States expand on evidence from the 2013 Pb ISA and
provide further support for an association between Pb exposure and health outcomes among minority
populations. Specifically, several analyses using NHANES data reported increases in BP among non-
Hispanic white individuals and non-Hispanic Black individuals (Appendix 4.3.1.1.1). However, increases
in BP and hypertension prevalence were consistently larger among non-Hispanic Black individuals. These
findings held true across several nationally representative cross-sectional studies. Taken together, the
evidence suggests that in addition to having higher BLLs, associations between blood Pb and BP and
hypertension are larger among non-Hispanic Black populations when compared with Hispanic or non-
Hispanic white populations. However, due to the cross-sectional nature of these studies, the observed
racial differences may also reflect a history of greater exposure to Pb among non-Hispanic Black
populations that is not fully captured in the concurrent BLL metric. Racial differences were also noted for
associations of Pb exposure and neurodevelopmental outcomes in children, but the evidence was limited
to a single study. Overall, recent evidence confirms and extends the previous ISA's findings, indicating
increases in Pb biomarker levels and a differential association between Pb exposure biomarkers and
changes in BP or hypertension status, and potentially neurodevelopmental outcomes in children based on
race/ethnicity.
IS.7.4.2.2 Childhood
Historically, children have been known to be at particularly higher risk for Pb-related health
effects. The 2013 Pb ISA provided a plethora of evidence indicating a greater likelihood of Pb-related
health outcomes among children. Previous toxicokinetic studies established that Pb can cross the placenta
and disrupt the developing nervous system of the fetus. Additionally, studies have shown that children's
behaviors and activities (including increased hand-to-mouth contact, pica behavior, crawling, and poor
handwashing), differences in diets (e.g., consumption of breast milk), and biokinetic factors may place
them at greater risk for exposure. There was strong evidence for Pb-related cognitive deficits and
behavioral problems across gestation, childhood, and into adolescence. Among adolescents, Pb exposure
was linked to delinquent or criminal behavior, delays in pubertal onset, and renal effects. However,
uncertainty exists regarding the timing and duration of Pb exposure on observed health effects because of
the high levels of Pb in the adolescent populations studied. Several studies reported evidence for Pb-
related increases in immunosuppression, immune sensitization, and allergic responses in children.
Associations were also found for increased anemia and reduced RBC function and survival. Children with
higher BLLs were also reported to be at higher risk for dental caries.
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Recent evidence extends support for Pb-related decrements in FSIQ, infant neurodevelopment,
learning, memory, executive function, and academic performance/achievement in children. Several recent
studies assessed timing of exposure by comparing associations between health outcomes and Pb levels
measured during different exposure windows, including during gestation, birth, early-life, and concurrent
exposures. There is no consistent pattern for critical exposure windows in the recent evidence base, which
is consistent with the heterogeneity of results observed for different timing and duration of exposures in
studies evaluated in the 2013 Pb ISA. Toxicological studies and epidemiologic studies examining
modification by genetic/epigenetic factors, coexposure to other metals, or maternal stress indicate that
time window sensitivities may be linked with biological, environmental, and psychosocial variables that
operate at different timepoints during development. A few studies provide evidence for the persistence of
effects of prenatal or early-life Pb exposure, noting early childhood cognitive deficits that continue into
late adolescence. Furthermore, two animal studies reported Pb-induced cognitive function effects with
longer exposure durations that spanned multiple developmental periods. Additionally, several
epidemiologic studies indicated nonlinear C-R relationships between BLLs and cognitive function in
children, which may be explained by unmeasured confounding or interaction by sex, genetics, underlying
conditions, sociodemographics, and timing or duration of exposure. Most studies, however, generally
supported dose-dependent cognitive function decrements at BLLs <30 (ig/dL.
Additional recent studies find strong evidence for Pb affecting externalizing behaviors in
children, including through influence on attentional deficits, impulsivity, hyperactivity, conduct disorders,
aggression, and criminal behavior. Similarly, gestational, postnatal, adolescent, and average childhood
(from birth to ages 4-5 or 11-13 years) Pb biomarker concentrations are associated with internalizing
behaviors, such as anxiety and depression. Both gross and fine motor function are also affected, in line
with previous findings involving oxidative stress, inflammation and Ca2+ signaling, impaired neuron
development, synaptic changes, and neurotransmitter changes with increased Pb exposure. No clearly
defined pattern exists regarding a specific sensitive exposure window regarding these health effects,
although a few toxicological studies report greater decrements in motor function in association with
gestational Pb exposure.
Although the 2013 Pb ISA found support for Pb-related immune effects in children, recent
evidence was less consistent. Results for immunosuppression are consistent with previous findings, but
the body of literature regarding immune sensitization and allergic responses was generally null. On the
other hand, recent evidence supports results from previous studies, reporting associations between Pb
exposures and decreased RBC survival and function, including increased prevalence of anemia among
children with mean BLLs <10 (ig/dL.
Recent epidemiologic studies also continue to report consistent associations between BLLs and
delayed puberty among male and female adolescents. Some studies suggest that as BLLs decline, the
association between blood Pb and age of menarche may be attenuated by potential confounders such as
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body weight and adiposity. Of note, there is some evidence that childhood BLLs may affect the function
of insulin-like growth factor, which could lead to delays in growth and pubertal onset in adolescent boys.
Although no recent toxicological studies have examined the relationship between Pb exposure
and teeth, several recent epidemiological studies among large populations reinforce previous conclusions
of increased dental caries in association with higher BLLs in early childhood.
Overall, substantial toxicokinetic, exposure, and health evidence continues to support the
previous conclusion that children are at increased risk for the health effects of Pb.
15.7.4.2.3 Proximity to Pb Sources
Studies from the 2013 Pb ISA provided sufficient evidence that living near Pb sources, including
large industrial sources and urbanized areas with Pb-contaminated soils, is associated with increased Pb
exposure. Additionally, aviation fuel was highlighted as a major source of Pb emissions in ambient air
(Appendix 1.2). A study in North Carolina reported inverse associations of children's BLLs with
proximity of their residence to airports (where leaded aviation fuel may be used). Recent evidence
continues to support increased Pb biomarker levels associated with proximity to airports and other Pb
sources. Additionally, recent evidence also implies that a reduction in environmental Pb at a particular
source (e.g., superfund site) is associated with decreases in the BLLs of children in proximity to the
original source.
In addition to increased biomarker Pb levels being associated with proximity to Pb sources,
several recent epidemiologic analyses have reported increased Pb-related health effects among those in
proximity to industrial sources of Pb. Specifically, studies comparing populations within certain distances
of a Pb source indicated increases in BP and decreases in renal function, though they did not control for
additional metals in their analyses. Additionally, recent studies have assessed child IQ and observed small
reductions in child intelligence in closer proximity to Pb sources.
15.7.4.2.4 Nutrition
The 2013 Pb ISA and prior AQCDs concluded that by limiting or outcompeting Pb for absorption
in the gastrointestinal tract, diets rich in minerals including Ca2+, Fe, and zinc give some protection from
increased BLLs. Additionally, previous epidemiologic and toxicological investigations indicated that
people with Fe deficits are at increased risk for Pb-related health consequences. Therefore, there are
sufficient data from several fields showing certain nutritional factors affect the risk of Pb exposure and
health effects in a population.
Recent epidemiologic studies continue to explore other modifications of Pb-related health effects
by diet or nutritional intake. An evaluation of the impact of two different diet types (Prudent: high
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amounts of fruit, legumes, whole grains, tomatoes, seafood, poultry, cruciferous vegetables, dark-yellow
vegetables, leafy vegetables, and other vegetables; Western: high intake of processed meat, red meat,
refined grains, butter, high-fat dairy products, eggs, and fries) was conducted on the relationship between
bone Pb levels and cardiovascular outcomes. In this study, patella Pb measurements among those with
low prudent diets were associated with a higher risk of coronary artery disease compared with those with
a high prudent diet. This difference was not evident when assessing tibia Pb measurements.
Recent toxicological studies investigated the impact of various dietary factors on the effects of Pb
on neurological outcomes. A recent study reported that in comparison with a standard diet, a high-fat diet
exacerbated the effect of Pb on learning deficits during the first stages of learning. Another study, which
supplemented Pb exposure in mice with green tea extract, reported that green tea ameliorated the negative
impact of Pb exposure on both learning and memory. Additionally, one study reported probiotic
supplementation partially mitigates the cognitive deficits observed in an active avoidance paradigm.
Given the disparate dietary factors examined across these studies, conclusions on the modifying potential
of any individual factor remains uncertain. However, when considered more generally, there is consistent
toxicological evidence that dietary factors modify the cognitive effects of Pb exposure.
Adding on to previous evidence from the 2013 Pb ISA, recent studies have connected Fe
deficiency to immune system effects in a few toxicological studies. A few studies that focused on
different outcomes reported decreases in anti-TT-specific IgM and mucosal IgA levels in rats that were
fed an Fe-deficient diet for 4 weeks and administered Pb acetate in drinking water for 4 weeks after
confirming Fe deficiency. Taken together, these studies support a role for dietary factors in the
immunotoxicity of Pb, but the diversity of nutritional factors investigated among a small number of
studies makes it difficult to determine their relative importance. In summary, the evidence continues to
indicate increased risk for populations with reduced intake of Ca2+ and Fe, and potential risk associated
with other dietary factors.
IS.7.4.2.5 Genetics
Evidence from the 2013 Pb ISA suggested that various genetic variants may modify the
relationship between Pb and various health effects. According to these previous epidemiologic and
toxicological studies, populations with specific ALAD variants may have increased risk of Pb-related
health effects. Variants of vitamin D receptor (VDR), dopamine receptor D4, glutathione S-transferase
(GST) Mu 1, tumor necrosis factor a, endothelial nitric oxide synthase, and the hemochromatosis gene
(HFE) were other genes studied, and presence of their variants may also affect the risk of Pb-related
health effects. Overall, the potential for genetic variants to modify Pb-related health outcomes were
investigated in a small number of studies. Therefore, despite some evidence that certain genetic variants
may modify Pb-related health effects, there are still some uncertainties in the evidence.
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Several recent studies in children add to the small body of previous evidence. The effect of Pb
exposure on children's IQ was reported to be weaker (i.e., smaller magnitude) among those with the
ALAD1 genotype (median BLL: 1.0 (ig/dL). This study identified unique glutamate ionotropic receptor
N methyl D aspartate-type subunit (GRIN)2A and GRIN2B variations that exacerbated Pb-related
deficiencies in learning, memory, and executive function, with a greater impact observed in boys.
Another recent study observed that prenatal Pb exposure was linked to DNA methylation in regions
including genes involved in neurodevelopment. Overall, there is limited evidence of interactions and
increased risk of relationships between genes and Pb exposure in children.
Several recent studies in adults have shown that certain genetic polymorphisms can be important
in assessing the potential for increased risk of Pb biomarker levels and of Pb-related health effects.
Specifically, VDR was evaluated in a longitudinal study examining the association between pulse
pressure (PP) and bone and BLLs. Variations in VDR genes have the potential to influence Pb
accumulation, absorption, and retention in the body. At the initial visit (baseline), an interquartile range
increase in either tibia or patella Pb was associated with an increased PP among those with the variant
(opposed to ancestral) genotype (single nucleotide polymorphisms [SNPs] in Bsml, Taql, Apal, or
Fokl). While the strength of the association between PP and tibia Pb diminished over time (10-year
follow-up), the three-way interaction terms between bone Pb, VDR receptor type, and time-since-baseline
was almost null, indicating that VDR consistently modifies the association between bone Pb and PP. In
another recent study, several other genes and proteins were also evaluated as effect measure modifiers of
the relationship between bone Pb measurements and incident CHD, including: ALAD, HFE, heme
oxygenase-1 (HMOX1), VDR, apolipoprotein E (APOE), GSTs, and renin-angiotensin. These genes and
the proteins they encode appear to play a role in influencing Pb uptake and retention, as well as altering
Pb toxicity. The authors constructed two sets of genetic risk scores summing either all of the measured
SNPs or a subset of SNPs that were observed to modify the relationship between Pb exposure and CHD.
The association between patella Pb levels and incident CHD was notably stronger in participants in the
highest tertiles of the two genetic risk scores compared with those in the lowest, suggesting that genetic
loci may modify Pb-related CHD risk.
Recent epidemiologic studies on gene regulation during pregnancy are limited but provide insight
on potential mechanistic pathways through which Pb may impact pregnancy. In one study, the association
between maternal Pb levels in blood, patella, and tibial bone and microRNA (miRNA) expression in the
cervix during the second trimester of pregnancy was assessed. Expression of two of the miRNAs were
associated with maternal second trimester BLLs. Another study assessed the association of BLLs during
pregnancy with mitochondrial DNA (mtDNA) content in cord blood, which is a sensitive marker of
mitochondrial function and oxidative stress. Maternal Pb levels during the second trimester were
associated with higher mtDNA content. As BLLs may differ by ALAD (aminolevulinic acid dehydratase)
genotype, one study compared growth outcomes in children with ALAD1-1 and ALAD1-2/2-2. There
were negative associations between baseline BLLs and height, knee height, and height-for-age z-score
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(HAZ). The observed associations between BLLs and height, knee height, and HAZ were stronger
(i.e., larger magnitude) in children with ALAD1-2/2-2 compared with ALAD1-1.
Overall recent studies have added to the body of evidence on genetic variants previously found to
modify the risk of Pb-related health effects. Recent studies have also identified other variants - including
but not limited to ALDA1, N-methyl D aspartate, HFE, VDR, HMOX1, and APOE - that may modify the
relationship of Pb exposure and human health effects and predispose certain populations to greater risk of
Pb-related health effects.
IS.7.4.2.6 Stress
The 2013 Pb ISA evaluated stress as a factor that could modify the association with Pb-related
health outcomes. Specifically, these effects were most commonly evaluated within studies evaluating
cognitive function. More recent evidence expands the knowledge base for stress as a factor that can
increase the risk of Pb-related health effects.
Several recent studies among children evaluated cardiovascular outcomes associated with Pb
biomarkers as a response to acute stressors. One study indicated that a higher level of Pb exposure during
early childhood (mean age of 2.6 years) was related to a greater total peripheral resistance response to
acute stress years later (at 9.5 years of age). Another study indicated significant decreases in HRV
associated with BLLs, following an acute stressful stimulus in young (aged 3-5) children.
Maternal stress has also been evaluated within studies assessing the relationship between
biomarkers of Pb exposure and neurologic and developmental outcomes among offspring. Maternal stress
appeared to substantially modify the associations between Pb exposure biomarkers and neurodevelopment
among children. Specifically, high maternal stress appeared to exacerbate the effect of prenatal Pb
exposure on neurodevelopment in several domains, including language. However, epidemiologic and
toxicological studies assessing birth outcomes did not observe an effect of maternal stress on relationships
between Pb and adverse birth outcomes. Overall, the majority of recent evidence strengthens the previous
conclusion that increased stress exacerbates the effects of Pb.
IS.8 Evaluation of Welfare Effects of Pb
Effects of Pb relevant to the secondary NAAQS are observed across ecological endpoints
common to terrestrial, freshwater, and saltwater biota. Those endpoints include reproduction, growth,
survival, neurobehavioral effects, hematological effects, and physiological stress, and occur at multiple
scales of biological organization from the cellular to the ecosystem. The atmosphere and terrestrial and
aquatic ecosystems are interconnected, with transfer of Pb taking place between each of these systems
(Appendix 11.1.2). Although Pb is present in the natural environment, it has no known biological function
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in plants or animals. In some instances, depending on the form of Pb and prevailing environmental
chemistry at a particular geographic location, Pb is taken up by biota where it can lead to a biological
response. Pb exposure of organisms can be via one or more pathways (e.g., uptake from soil or water,
ingestion). For Pb to interact with a biological membrane and be taken up into an organism it must be
bioavailable (Appendix 11.1.6). Generally, the greater amount of Pb available as the free Pb ion, the
greater the bioavailability. Factors such as pH, dissolved organic carbon (DOC) or water hardness in
aquatic environments, and pH, cation exchange capacity (CEC), or aging in terrestrial environments often
interact strongly with Pb concentration to modify its effects, primarily through their influence on
bioavailability, but also sometimes through direct modification of biotic effects. Uptake, subsequent
bioaccumulation, and toxicity of Pb varies greatly between species and across taxa, as characterized in the
1977 AQCD (U.S. EPA. 1986b). the 2006 Pb AQCD (U.S. EPA. 2006). the 2013 Pb ISA (U.S. EPA.
2013a). and further supported in this ISA. In natural environments it is difficult to attribute observed
effects solely to Pb due to the presence of confounding factors such as other pollutants, and additional
modifying factors that affect Pb bioavailability and toxicity. Furthermore, the portion of Pb from
atmospheric sources is usually not known. The welfare effects of Pb summarized in the following sections
are presented in greater detail in Appendix 11. Effects of Lead in Terrestrial and Aquatic Ecosystems.
Appendix 11 includes an overview of concepts related to ecosystem effects of Pb (Appendix 11.1) and
evidence for effects of Pb on organisms inhabiting terrestrial (Appendix 11.2), freshwater
(Appendix 11.3) and saltwater (Appendix 11.4) environments, especially since the 2013 Pb ISA.
Initial perturbations associated with exposure to Pb such as cytological or biochemical changes
may lead to effects at higher levels of biological organization (i.e., from the subcellular and cellular level
through the individual organism and up to ecosystem-level processes). The alteration of cellular ion status
(including disruption of Ca2+ homeostasis, altered ion transport mechanisms, and perturbed protein
function through displacement of metal cofactors) appears to be the major unifying mode of action
underlying all subsequent modes of action of Pb toxicity in plants, animals, and humans (Lassiter et al..
2015; U.S. EPA. 2013a). Molecular mechanisms linked to oxidative stress may induce DNA damage and
generation of reactive oxygen species leading to protein modification, lipid peroxidation, and altered
enzyme response. For ecological endpoints in this ISA, biochemical (e.g., enzymes, stress markers)
responses at the suborganism-level of biological organization are grouped under the broad endpoint of
"physiological stress," while organism-level effects include reproduction, growth, and survival. These
endpoints in turn have the potential to alter population, community, and ecosystem levels of biological
organization (Suter et al.. 2004). The definition of an ecosystem used in this ISA is "a functional unit
consisting of living organisms, their nonliving environment, and the interactions within and between
them" (Allwood et al.. 2014). Ecosystems can be natural, cultivated, or urban (U.S. EPA. 1986b) and may
be defined on a functional or structural basis (Appendix 11.1.4). Ecosystem structure includes species
abundance, richness, distribution, diversity, evenness, and composition measured at the population, or,
community scales, which may be further defined by spatial boundaries such as those relating to an
ecosystem, region, or global scale. Pollutants, such as Pb, can affect the ecosystem structure at any of
these scales, corresponding to levels of biological organization (Suter et al.. 2005). Causality
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determinations for ecological effects of Pb in this ISA use biological scale as an organizing principle to
summarize effects on vegetation, invertebrates, and vertebrates in terrestrial, freshwater, and saltwater
environments.
IS.8.1 Summary of Effects on Terrestrial Ecosystems
In terrestrial ecosystems, non-air media can receive Pb from atmospheric deposition or other
sources. Once deposited, Pb can be resuspended into the air or transferred among other environmental
media (Appendix 1.3). Since the 2013 Pb ISA (U.S. EPA. 2013a). evidence has continued to accrue for
many of the effects of Pb on terrestrial ecosystems reported in that ISA and previous U.S. EPA
assessments. In particular, effects previously documented were observed at exposures lower than in
previous studies. This additional supporting evidence includes investigations of effects on species and
communities that had not been previously studied, but the additional evidence is not sufficient to change
any of the causality determinations for terrestrial ecosystems that were reached in the 2013 Pb ISA.
Studies published since the 2013 Pb ISA (U.S. EPA. 2013a) continue to support previous findings
that plants generally sequester larger amounts of Pb in roots relative to shoots and that there are species,
ecotype, and cultivar-dependent differences in the uptake of Pb from soil and the atmosphere, as well as
in the translocation of sequestered Pb (Appendix 11.2.1). In the 2013 Pb ISA and previous assessments,
Pb exposure was found to result in plant physiological stress and deficits in plant growth, whereas
evidence of effects on plant survival and reproduction was mixed. Recent studies have continued to
demonstrate various deleterious physiological effects of Pb exposure on plants, particularly oxidative
stress. Strong uncertainties also remain regarding the concentrations at which these effects would be
observed in the environment. Recent studies have examined the protective effects of mycorrhizae and of
some plant nutrients when added in excess of the minimal requirements of the plants.
In terrestrial invertebrates, the 2013 Pb ISA (U.S. EPA. 2013a) and previous assessments
reported evidence of effects on invertebrates that included responses of antioxidants, reductions in growth
and survival, as well as decreased fecundity. Neurobehavioral aberrations and endocrine impacts were
also found, as well as incomplete evidence of hematological effects. Second-generation effects were also
observed. Evidence published since then provides additional support for the effects of Pb exposure on
organismal and suborganismal responses including a decrease in survival as well as decreased growth and
fecundity (Appendix 11.2.4.3). Recently published studies on physiological responses to Pb include
decreases in protein and lipid content and increases in malondialdehyde in earthworms.
Acetylcholinesterase activity decreased in response to Pb in snails and honeybees while the effects on
protein, glycogen, other enzymes, and GST responses were variable depending on the site or species
examined. Several new studies quantified changes in feeding and foraging behavior in bees following Pb
exposure. Evidence also suggests that in earthworms, Pb exposure can have lasting effects on growth
even postexposure and slow the time to maturation. Pb exposure delayed onset of the breeding season and
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shortened duration in isopods, as well as influenced mate selection in fruit flies. Evidence published after
the 2013 Pb ISA (U.S. EPA. 2013a) includes new organisms as well as modifying factors of organism
response such as habitat, exposure history, and seasonality.
Effects of Pb observed in terrestrial vertebrates include decreased survival and reproduction, as
well as neuro-behavioral effects and effects on development (U.S. EPA. 2006). The 2013 Pb ISA (U.S.
EPA. 2013a) also provided evidence for Pb effects on hormones, blood, and other physiological and
biochemical variables (U.S. EPA. 2013a). Evidence of effects on growth was limited. Studies published
since the 2013 Pb ISA provide additional evidence for effects on suborganism- and organism-level
endpoints, and specifically on hematological and physiological endpoints (Appendix 11.2.4.4). New
studies have expanded upon the relationship between Pb exposure and ALAD activity by adding more
species of birds, amphibians, and mammals to the evidence base. Additional evidence of oxidative stress
has been gathered, as well as evidence of effects on corticosterone levels and immunity in birds. Recent
literature continues to add to evidence relating to reproductive effects at both the organism and
suborganism levels including effects on lifetime breeding success and some specific secondary sexual
traits. New findings of behavioral effects of Pb included increased aggression in mockingbirds.
Systematic studies of the validity of using results of experiments with addition of soluble salts of
Pb to soil for estimating effects of Pb exposure under field conditions have continued since the 2013 Pb
ISA. As in previous work, recent experiments showed that the form of Pb, pH, CEC, organic carbon, Fe
and Mn oxides, percolation, aging, and soil composition are all strong modifiers of toxicity. Recent
studies demonstrated additional interactions among those variables and showed that their effects are at
times mediated by additional variables, such as salinity. Those studies add support to the conclusion that
data from exposure-response experiments in terrestrial environments conducted using spiking of soils
with soluble salts of Pb are unlikely to generate accurate estimates of effects in contaminated natural
environments (Appendix 11.2.5). However, Ports et al. (2021) suggested that two corrections to the
results of exposure-response experiments conducted with additions of soluble salts of Pb to soil may be
sufficient to derive predicted no-effect concentrations according to the European Registration, Evaluation,
Authorisation and Restriction of Chemicals Regulation (European Parliament and Council. 2006).
According to the 2013 Pb ISA (U.S. EPA. 2013a) and previous assessments, effects on terrestrial
communities and ecosystems observed in contaminated natural environments have included decreased
species diversity, changes in floral and faunal community composition, and decreasing vigor of terrestrial
vegetation. In addition to impacts on soil microbial community function alone, interconnection of effects
of Pb contamination among soil bacterial and fungal community structure, earthworms, and plant growth,
have also been systematically documented. Some new evidence of the effects of Pb at the terrestrial
community and ecosystem levels of biological organization has since been added. Many studies on the
effects of Pb on microbial communities were reported in the 2013 Pb ISA (U.S. EPA. 2013a). Additional
observational studies published since then (Appendix 11.2.4.1). many of which were anthropogenic
environmental gradient studies, have continued linking Pb exposure and effects on microbial community
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structure (e.g., abundance, diversity) and function (e.g., enzyme activities, respiration rates). Many found
mixed (negative, positive, or null) relationships between total or bioavailable Pb soil concentration and
the abundance of bacterial and fungal taxa. It remains difficult to disentangle the effects of Pb exposure
on microbial communities from the effects of other soil contaminants using anthropogenic environmental
gradients, as other heavy metals and soil physicochemical properties are significantly correlated with soil
Pb concentration, and many of these factors also influence microbial processes. In addition to microbial
communities, species interactions between tree species and their pests, and between herbaceous plants
and nectar robbers, worms, and lepidopteran consumers were among the new community and ecosystem
endpoints for which effects of Pb were observed (Appendix 11.2.6). Several studies found inverse
relationships between Pb concentration along a pollution gradient and community structure of soil mites,
potworms, nematodes, and invertebrates associated with kale plants. Although evidence for effects on
growth, reproduction, and survival at the individual organism level and in simple trophic interactions
makes the existence of effects at higher levels of organization likely, direct evidence is relatively sparse
and difficult to quantify. The presence of multiple stressors, especially other metals, continues to
introduce uncertainties in attributing causality to Pb at higher levels of organization.
IS.8.2 Summary of Effects on Freshwater Ecosystems
Freshwater organisms including algae, aquatic plants, microbes, invertebrates, vertebrates, and
other biota with an aquatic lifestage (e.g., amphibians) may be exposed to Pb in aquatic environments.
Inputs of Pb to freshwater ecosystems include air-related sources and non-air sources. Atmospherically
derived Pb can enter aquatic systems through direct wet or dry deposition and erosional transport or
resuspension of Pb from terrestrial systems (Appendix 11.1.2). Receiving water bodies include lakes
(lentic systems) and rivers and streams (lotic systems). Freshwater wetlands, some of which may be
inundated occasionally or constantly, also provide habitat for aquatic biota. Uptake of Pb by aquatic biota
may occur via multiple exposure routes including direct absorption from the water column, ingestion of
contaminated food and water, uptake from sediment porewater, or incidental ingestion of sediment (U.S.
EPA. 2013a. 2006).
As described in previous U.S. EPA reviews of Pb, sensitivity to this metal can vary by several
orders of magnitude across freshwater biota. Pb elicits responses in some freshwater invertebrate species
at concentrations below 5 to 10 (ig Pb/L (under some water conditions) while other freshwater organisms
appear to be unaffected at concentrations greatly exceeding 1,000 |ig Pb/L. Most of the available studies
of Pb exposures in freshwater biota are laboratory toxicity tests on single species in which an organism is
exposed to a known concentration of Pb, and the effect on a specific endpoint is evaluated. Concentration-
response data from freshwater organisms indicate that there is a gradient of response to increasing Pb
concentration and that some effects in sensitive species are observed at or near the upper limit of Pb
concentrations quantified in U.S. surface waters (Appendix 11. Table 11-1). Freshwater invertebrate taxa
that exhibit sensitivity to Pb include some species of gastropods, amphipods, cladocerans, and rotifers,
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although the toxicity of Pb is highly dependent upon water quality variables such as DOC, hardness,
and pH.
Physicochemical properties of surface waters such as hardness, DOC, and pH can be quantified,
are directly related to the toxic effects, and are used in bioavailability models to predict acute and chronic
toxicity (Appendix 11.1.6). As described in prior AQCDs, the 2013 Pb ISA, and this document
(Appendix 11.3.2.1.1). the effect of water hardness is variable; generally, both the acute and chronic
toxicity of Pb increases with decreasing water hardness as Pb becomes more soluble and bioavailable and
less Ca2+ and Mg2+ ions are available to compete with Pb for binding sites. DOC has a protective effect on
Pb toxicity in freshwater invertebrates and fish; newer studies generally continue to support these
observations with some exceptions (Appendix 11.3.2.1.2). Since the 2013 Pb ISA, studies have further
elucidated the relationship between the characteristics of humic substances and Pb bioavailability. As
described in prior AQCDs and the 2013 Pb ISA, uptake and subsequent toxicity of Pb to freshwater biota
can also be affected by pH, either directly or indirectly (Appendix 11.3.2.1.3). Generally, at low pH, there
is more Pb2+ available to bind to the biotic ligand. As pH increases, there is increased formation of Pb
organic (DOC) and inorganic (OH-, COr ) complexes, which decrease Pb bioavailability. Since the 2013
Pb ISA, several studies have further characterized Pb complexation and adsorption under changing pH
conditions, recent studies generally support the previous understanding that higher pH is protective; these
findings vary by the duration of the toxicity bioassays and by taxa, however.
Biological factors that may influence freshwater organism response to Pb exposure include
lifestage, genetics, and nutrition (see Section 7.2.3, 2006 Pb AQCD, Section 6.4.9, 2013 Pb ISA, and
Appendix 11.3.2 of this ISA). These factors are more difficult to link quantitatively to toxicity than water
chemistry variables. Often, species" differences in metabolism, sequestration, and elimination rates
influence the relative sensitivity and vulnerability of exposed organisms. Uptake studies generally show
that aquatic invertebrates and vertebrates accumulate Pb from water in a concentration-dependent manner
and may reach an equilibrium depending on the organism's ability to eliminate or store Pb. Since the
2013 Pb ISA, several studies have examined how the activities of sediment-associated benthic
invertebrates (sometimes called ""bioturbators" because of the biological role they play in water column
turbidity) influence Pb transfer to the water column and subsequent bioavailability to other aquatic
organisms (Appendix 11.3.2.1.11). Overall, presence of these bioturbators can enhance Pb availability to
organisms in the water column and potentially cause toxic effects in those organisms.
For freshwater plants and algae, studies on bioavailability and toxicity of Pb published since the
2013 Pb ISA (Appendix sections 11.3.2.2 and 11.3.4.2) continue to support previous findings that plants
tend to sequester larger amounts of Pb in roots as compared with shoots and that there are species-specific
differences in uptake of Pb, compartmentalization of that sequestered Pb, and plant response (U.S. EPA.
2013a. 2006). Most studies on effects of Pb in freshwater algal species reviewed in the 2013 Pb ISA and
the AQCDs were conducted with nominal media exposures and effect concentrations greatly exceeded Pb
reported in surface water. In the 1977 Pb AQCD, differences in sensitivity to Pb among different species
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of algae were observed, and concentrations of Pb within the algae varied among genera and within a
genus (U.S. EPA. 1977). The 1986 Pb AQCD (U.S. EPA. 1986b) reported that some algal species
(e.g., Scenedesmus sp.) were found to exhibit physiological changes when exposed to high Pb
concentrations in situ. Effects of Pb on algae reported in the 2006 Pb AQCD included decreased growth,
deformation, and disintegration of algae cells, and blocking of the pathways that lead to pigment
synthesis, thus affecting photosynthesis. New information since the 2013 Pb ISA includes studies on
common reed (Phragmites australis) showing significant decreases in total biomass, photosynthesis, and
rhizome growth as well as alterations in growth form and propagation strategy under Pb exposure and a
study in freshwater algae based on analytically verified concentration of Pb (Appendix 11.3.4.2 and
Table 11-5).
Freshwater aquatic invertebrates are generally more sensitive to Pb exposure than other taxa.
Controlled studies at concentrations near the upper range of representative concentrations of Pb available
from surveys of U.S. surface waters (median: 0.50 |ig Pb/L; range 0.04 to 30 |ig Pb/L, 95th percentile
1.1 Mg Pb/L) (U.S. EPA. 2006). reviewed in the 1986 AQCD, the 2006 Pb AQCD, the 2013 Pb ISA, and
this document provide evidence for the effects of Pb on reproduction, growth and survival in sensitive
freshwater invertebrates, notably gastropods, cladocerans, rotifers, and amphipods. In studies reviewed in
the 2013 Pb ISA the freshwater snail (Lymnaea stagnalis) was identified as one of the most sensitive
species to Pb exposure, and more recent studies support these observations. Recent evidence further
characterizes Pb effects on growth and reproduction at concentrations below 10 |ig Pb/L in sensitive
species of freshwater gastropods, cladocerans, rotifers, and amphipods, especially under chronic exposure
scenarios (Appendix 11.3.5 and Table 11-5).
For freshwater vertebrates, early studies on waterfowl investigated exposure to Pb via accidental
poisoning or ingestion of Pb shot (U.S. EPA. 1977). Studies on aquatic vertebrates reviewed in the 1986
Pb AQCD were limited to hematological, neurological, and developmental responses in fish (U.S. EPA.
1986b). In the 2006 Pb AQCD, effects on freshwater vertebrates included consideration of the role of
water quality parameters on toxicity to fish, as well as limited information on the sensitivity of turtles and
aquatic stages of frogs to Pb (U.S. EPA. 2006). Evidence in the 2013 Pb ISA supported the 2006 Pb
AQCD conclusions that the gill is a major site of Pb uptake in fish and that there are species differences in
the rate of Pb accumulation and distribution of Pb within the organism. Several studies in fish in which Pb
concentration was analytically verified provide additional evidence for reproductive and developmental
effects for freshwater vertebrates (Appendix 11.3.4.4.1.2). New studies continue to show distinct patterns
of Pb tissue distribution in water versus dietary exposures (Appendix 11.3.2.4).
Reductions in species abundance, richness, and diversity associated with the presence of Pb in
freshwater habitats are reported in the literature, usually in heavily contaminated sites where Pb (and
other metal) concentrations are higher than typically observed environmental concentrations. Most
evidence is from sediment-associated macroinvertebrate communities. New studies generally confirm
findings in the 2006 Pb AQCD (U.S. EPA. 2006) and 2013 Pb ISA (U.S. EPA. 2013a) that transfer of Pb
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through the food web is generally low (Appendix 11.3.2.5). Observational and experimental studies
published since the 2013 Pb ISA continue to show negative associations between sediment and/or
porewater Pb concentration and macroinvertebrate communities (Appendix 11.3.6). The evidence is
expanded somewhat with studies reporting associations with Pb and periphyton abundance.
Approaches for characterizing the toxicity of Pb to freshwater biota since the 2013 Pb ISA
include a proposal for updating aquatic life water quality criteria (Appendix 11.1.7.3). The existing
U.S. EPA ambient water quality criteria (AWQC) for Pb for the protection of aquatic life are a criterion
maximum concentration of 65 |ig Pb/L (for acute exposure) and criterion continuous concentration of
2.5 |ig Pb/L (for chronic exposure) at a hardness of 100 mg/L (U.S. EPA. 1985). Using the biotic ligand
model (BLM) (Appendix 11.1.6) (Deforest et al.. 2017) proposed acute BLM-based freshwater criteria
ranging from 18.9 to 998 |ig Pb/L and chronic BLM-based Pb freshwater criteria ranging from 0.37 to
41 jag Pb/L. The lowest criteria were for water with low DOC (1.2 mg/L), pH (6.7) and hardness
(4.3 mg/L as CaCOs), and the highest criteria were for water with high DOC (9.8 mg/L), pH (8.2) and
hardness (288 mg/L as CaCOs), which encompasses varying water quality conditions of North American
surface waters. The updated data sets in Deforest et al. (2017) incorporated toxicity information for L.
stagnalis, the cladoceran, Ceriodaphnia dubia, the amphipod, Hycdella azteca, and the rotifer, Philodina
rapida; freshwater invertebrates that are relatively sensitive to Pb exposure. Compared to the number of
genera used to develop the existing U.S. EPA AWQC for Pb (1984) for the protection of aquatic life, the
number of genera with acute toxicity data for Pb increased from 10 to 32, and for chronic toxicity, from 4
to 13, which enabled the proposed chronic criteria to be based on bioassay data rather than an acute-to-
chronic ratio. Additional advances in freshwater research since the 2013 Pb ISA have included
development and evaluation of bioavailability models to predict the toxicity of acute and chronic metal
mixtures, of which Pb is one component (Appendix 11.3.2.1.5). Considerable research beyond the scope
of this document (Appendix 11.1.1) has focused on metal mixture assessment, including how uptake and
bioaccumulation are affected in freshwater biota in the presence of multiple metals.
IS.8.3 Summary of Effects on Saltwater Ecosystems
Saltwater ecosystems encompass a range of salinities from just above that of freshwater (<1 ppt)
to that of seawater (generally described as 35 ppt). These ecosystems may receive Pb from multiple
sources such as contributions from direct atmospheric deposition and via inputs from terrestrial systems
including runoff and riverine transport (Appendix 1). Habitats associated with coastal areas include salt
marshes, estuaries, shallow open waters, sandy beaches, mud and sand flats, rocky shores, oyster beds,
coral reefs, mangrove forests, river deltas, tidal pools, and seagrass beds (U.S. EPA. 2016b). Estuaries,
where freshwater inflows gradually mix with salt water, are dynamic, heterogeneous environments
characterized by physicochemical gradients of salinity. The Pb2+ ion, which is the most bioavailable form
of Pb, is not common in seawater; rather, Pb primarily exists as a carbonate complex and to a lesser extent
as a chloride complex (Appendix 11.4.1).
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Factors affecting bioavailability of Pb to saltwater organisms (Appendix 11.4.2) are many of the
same factors affecting bioavailability to freshwater biota, notably OM and pH. Since the 2013 Pb ISA,
studies have further explored the effects of varying dissolved OM composition and changing pH on Pb
uptake and bioaccumulation in saltwater biota. In contrast to freshwater, OM in saltwater systems does
not necessarily demonstrate a protective effect and in some cases exacerbated toxicity of Pb to
invertebrates (Appendix 11.4.2.1). Other factors, such as salinity, play a greater role in Pb fate, transport,
and bioavailability in marine and estuarine systems, especially in dynamic estuarine environments
characterized by physicochemical gradients of salinity (Appendix 11.4.2.3). Other factors that affect
uptake and toxicity of Pb, such as biological adaptations by organisms, and the role of seasonality,
metabolism, diet, and lifestage, are more difficult to link quantitatively to toxicity (Appendix 11.4.2).
For saltwater plants, there is relatively little information on biouptake and toxicity at
concentrations of Pb typically encountered in the environment. Limited data on marine algae and
saltwater plants reviewed in the 1986 Pb AQCD, 2006 Pb AQCD, the 2013 Pb ISA and a few new studies
(Appendix 11. sections 11.4.2.10 and 11.4.4.2) provide evidence for species differences in Pb uptake,
bioaccumulation rates and toxicity. As in freshwater plants, Pb is concentrated in root tissue, but
sensitivity is species specific. Understanding of Pb effects in saltwater plants has not changed appreciably
since the 2013 Pb ISA; observed effects occur at much higher Pb exposures than are found in the natural
environment.
The majority of available studies of Pb effects on saltwater organisms are for invertebrate species.
Uptake and subsequent bioaccumulation of Pb in marine invertebrates varies greatly between species and
across taxa (U.S. EPA. 2006) (U.S. EPA. 2013a) and Appendix 11.4.2.11. In the 2006 Pb AQCD, a few
effects were noted in saltwater invertebrates including differences in sensitivity to Pb in copepods,
increasing toxicity of Pb with decreasing salinity in mysids, and effects on embryogenesis in bivalves
(U.S. EPA. 2006). In the 2013 Pb ISA, several studies reported concentrations associated with
reproductive effects in saltwater invertebrates including in a marine amphipod, a polychaete, and clams
(U.S. EPA. 2013a). Several field monitoring studies with marine bivalves in the 2013 Pb ISA used ALAD
as a biomarker for Pb exposure and correlated ALAD inhibition to increased Pb tissue content. Field and
laboratory studies provide evidence for antioxidant response to Pb exposure; however, most effects are
observed at concentrations of Pb that are higher than concentrations detected in marine environments.
New information for saltwater invertebrates since the 2013 Pb ISA includes additional studies that report
physiological perturbations associated with Pb exposure, including a few observations in previously
untested taxa. Recent exposure-response data for saltwater invertebrates (Appendix 11.4.5 and
Table 11-7) include reproductive and developmental bioassay results based on analytically verified
concentration for mollusks and echinoderms, with effects reported at lower concentrations than studies
included in the 2013 Pb ISA. Specifically, several embryo development bioassays for bivalves (48-hr
exposure) and sea urchin (72-hr exposure) found effects at concentrations <50 |ig Pb/L with no effects at
concentrations <10 (ig Pb/L for a few species (Markich. 2021; Romero-Murillo et al.. 2018; Nadella et al..
2013).
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For saltwater vertebrates, available information is largely for fish, with a few field-based studies
in birds and sea turtles (Appendix 11.4.4.4). Studies published since the 2013 Pb ISA provide chronic
toxicity data for several fish species, information that was previously lacking for evaluating longer-term
effects of Pb on these organisms. Calculated chronic no-observed-effect concentrations (NOECs) for
three saltwater fish species are <15 |ig Pb/L with effects reported in the range of 15 to 30 (ig Pb/L for
survival (Appendix 11. Table 11-7). These studies in fish were conducted with juvenile lifestages.
For community- and ecosystem-level effects evidence from field studies in saltwater
environments in the 2006 Pb AQCD and the 2013 Pb ISA, studies found either negative or null
relationships between Pb and species abundance, richness, and diversity in saltwater macroinvertebrates;
Pb is not the only contaminant in most observational studies, however, thereby making it difficult to
separate the effects of Pb alone from other metal pollutants. Several experimental and observational
studies since the 2013 Pb ISA reported negative relationships between sediment or saltwater Pb
concentration and microbial abundance and diversity, while other studies found no relationship
(Appendix 11.4.4.1). Additionally, some studies since the 2013 Pb ISA find reductions in foraminiferal
and/or meiofaunal community richness, diversity, and/or abundance associated with higher concentrations
of Pb in sediment and water, while others find positive or null correlations (Appendix 11.4.6). New
observational studies in saltwater systems generally confirm findings in the 2006 Pb AQCD (U.S. EPA.
2006) and 2013 Pb ISA (U.S. EPA. 2013a) of little transfer of Pb through the food web, with Pb
concentration decreasing with increasing trophic level (Appendix 11.4.2.13).
Since the 2013 Pb ISA, there are new toxicity data for saltwater biota that address some of the
uncertainties at that time. There are new studies reporting effects of Pb on survival in saltwater
vertebrates (Appendix 11.4.5) and additional evidence for reproductive and developmental effects in
saltwater invertebrates (Appendix 11.4.5). Furthermore, in many of the studies supporting these effects,
the concentration of Pb in the exposure media is analytically verified. This information reduces
uncertainties identified in the previous review concerning a lack of exposure-response data for saltwater
organisms, especially for chronic toxicity, and enables calculations of effect levels for saltwater biota
based on experimental data. An increase in toxicological data for saltwater organisms over the last several
years and availability of studies that analytically verify Pb exposure concentration has led to a study
proposing updates to the acute and chronic AWQC for Pb (Church et al.. 2017). For the acute criterion,
the newly proposed value of 100 |ig Pb/L is less than the current acute criterion of 210 (ig Pb/L due to
more recent acute toxicity data from relatively sensitive early lifestages of echinodermata and mollusca.
The proposed chronic criterion for saltwater biota is 10 |Lig Pb/L. Finally, there is additional evidence for
Pb association with changes in benthic invertebrate, microbial, and foraminiferal communities in coastal
environments (Appendix 11.4.4.1 and 11.4.6).
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IS.8.4
Summary of Welfare Effects Evidence
In the 2013 Pb ISA (U.S. EPA. 2013a). a series of causality determinations were made for effects
of Pb on plants, invertebrates, and vertebrates in terrestrial, freshwater, and saltwater ecosystems (U.S.
EPA. 2013a). Evidence published since that time supports or slightly expands the evidence for endpoints
that were already established as causal in the 2013 Pb ISA (Table IS-14). A few studies report effects at
lower effect concentration than in the 2013 Pb ISA. The new evidence is not sufficient to change any of
the previous causality determinations for terrestrial and freshwater organisms and ecosystems. New
evidence for terrestrial (Appendix 11-2) and freshwater (Appendix 11-3) biota continue to support
the existing causality determinations from the 2013 Pb ISA summarized in Table IS-14.
At the time of the 2013 Pb ISA there were fewer studies on effects of Pb in saltwater biota
compared with terrestrial and freshwater organisms and evidence was inadequate to infer causality
relationships for many endpoints. Specifically, there was a lack of chronic toxicity data, and relatively
few studies reported analytically verified Pb concentration in the experimental media. Several newer
studies quantify Pb in water and/or sediment and report effects on endpoints at lower concentration than
previously observed for saltwater biota, some of these studies are chronic exposure bioassays. Since the
2013 Pb ISA, the additional research for saltwater organisms supports a change in causality
determinations for three endpoints (Table IS-14). Specifically, the evidence is sufficient to conclude
there is likely to be a causal relationship between Pb exposure and reproductive and developmental
effects in saltwater invertebrates. In addition, the evidence is suggestive of, but not sufficient to infer,
a causal relationship between Pb exposure and saltwater vertebrate survival, and, the evidence is
suggestive of, but not sufficient to infer, a causal relationship between Pb exposure and saltwater
community and ecosystem effects. Previous causality determinations for the remaining saltwater
endpoints shown in Table IS-14 remain unchanged from the 2013 Pb ISA.
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Table IS-14 Summary of causality determinations for welfare effects of Pb
Level
Effect
Terrestrial3
Freshwater3
Saltwater3
Community-
and
Ecosystem
Community and Ecosystem Effects
Likely Causal
Likely Causal
TSuggestive
Reproductive and Developmental Effects - Plants
Inadequate
Inadequate
Inadequate
(0
+-»
C
o
C
o
Q_
Reproductive and Developmental Effects -
Invertebrates
Causal
Causal
TLikely
Causal
Q.
¦o
c
LD
Reproductive and Developmental Effects -
Vertebrates
Causal
Causal
Inadequate
Growth - Invertebrates
Likely Causal
Causal
Inadequate
TO
_i
i
E
w
"c
Growth - Vertebrates
Inadequate
Inadequate
Inadequate
Q.
O
Survival - Plants
Inadequate
Inadequate
Inadequate
03
S?
o
Survival - Invertebrates
Causal
Causal
Inadequate
Survival - Vertebrates
Likely Causal
Causal
TSuggestive
Neurobehavioral Effects - Invertebrates
Likely Causal
Likely Causal
Inadequate
Neurobehavioral Effects - Vertebrates
Likely Causal
Likely Causal
Inadequate
Hematological Effects - Invertebrates
Inadequate
Likely Causal
Suggestive
E w
.2 S>
Hematological Effects - Vertebrates
Causal
Causal
Inadequate
03 O
U) Q.
Physiological Stress - Plants
Causal
Likely Causal
Inadequate
O
-------
physiological stress in terrestrial, freshwater, or saltwater biota (Table IS-15) At the time of the 2013 Pb
ISA, evidence was sufficient to conclude that there is a causal relationship between Pb exposures and
physiological stress in terrestrial plants, and new evidence has reinforced this conclusion. Evidence is
sufficient to conclude that there is a likely to be causal relationship between Pb exposures and
physiological stress in terrestrial invertebrates and vertebrates as well as freshwater plants, invertebrates,
and vertebrates. Further evidence in saltwater invertebrates is suggestive of a causal relationship between
Pb exposures and physiological stress. Evidence is inadequate to conclude that there is a causal
relationship between Pb exposure and physiological stress responses in saltwater plants and vertebrates.
Recent literature supports the previous evidence for Pb effects on enzymes and antioxidant activity in
freshwater invertebrates (Appendix 11.3.4.3.1). New studies on physiological stress endpoints in
freshwater invertebrates include changes in the activities of antioxidant defense enzymes such as
superoxide dismutase, catalase, and glutathione peroxidase with aqueous exposure to Pb. A large body of
evidence supports sublethal biomarker perturbations with Pb exposure in freshwater vertebrates; however,
few studies were identified for this ISA that reported physiological response at more environmentally
relevant concentrations of Pb (<10 (ig Pb/L; Appendix 11.1.1) or concurrently assessed response at
organism-level endpoints (i.e., from the cellular and subcellular level to effects on growth, reproduction,
or survival).
Table IS-15 Summary of evidence for effects of Pb on physiological stress
endpoints in terrestrial and aquatic biota
Evidence from the 2013 Pb ISA Evidence from the 2024 Pb ISA
Terrestrial Plant Physiological Stress: Causal
Several studies from the 2006 Pb AQCD report lipid
peroxidation in plants; however, exposures in these studies
were higher than would be found generally in the
environment (U.S. EPA. 2006). Building on the body of
evidence presented in the 2006 Pb AQCD, studies in the
2013 Pb ISA provide evidence of upregulation of antioxidant
enzymes and increased lipid peroxidation associated with
Pb exposure in many species of plants. Increased
antioxidant enzymes with Pb exposure occur in some
terrestrial plant species at concentrations approaching the
average Pb concentrations in U.S. soils.
Freshwater Plant Physiological Stress: Likely to Be Causal
Increases of antioxidant enzymes with Pb exposure occur in
algae, mosses, and floating and rooted aquatic
macrophytes. Most available evidence for antioxidant
responses in aquatic plants are from laboratory studies
lasting from 2 to 7 d and at concentrations higher than
typically found in the environment. However, data from
transplantation experiments from nonpolluted to polluted
sites indicate that elevated enzyme activities are associated
with Pb levels measured in sediments.
Saltwater Plant Physiological Stress: Inadequate
Insufficient evidence to assess causality Insufficient evidence to assess causality
Recent studies continue to confirm increased
antioxidant activity in plants in response to Pb stress
as well as genotoxic effects of Pb exposure
(Appendix 11.2.4.2).
Physiological stress response in freshwater
vegetation is typically observed at much higher Pb
exposures than are found in the natural environment.
Studies reporting antioxidant processes upregulated
in algae support previous findings of a likely to be
causal relationship (Appendix 11.3.4.2).
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Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Terrestrial Invertebrate Physiological Stress: Likely to Be Causal
Changes in enzyme activities and oxidative stress markers
were reported in terrestrial invertebrates, including
earthworms, snails, and nematodes.
Additional studies in a few terrestrial invertebrate
species, notably earthworms, report altered enzyme
activity and perturbations in other biomarkers of
physiological stress associated with Pb exposure
(Appendix 11.2.4.3.1).
Freshwater Invertebrate Physiological Stress: Likely to Be Causal
Stress responses associated with exposure to Pb in aquatic
invertebrates reported in previous AQCDs include elevated
heat shock proteins, osmotic stress, lowered metabolism,
and decreased glycogen levels. Although these stress
responses are correlated with Pb exposure, they are
nonspecific and may be altered with exposure to any
number of environmental stressors. Evidence in the 2013 Pb
ISA included upregulation of antioxidant enzymes,
production of reactive oxygen species, and increased lipid
peroxidation associated with Pb exposure.
Recent literature (Appendix 11.3.4.3.1) supports
previous findings of Pb effects on enzymes and
antioxidant activity in freshwater invertebrates.
Physiological stress response was also observed in
several invertebrates in chronic sediment bioassays
conducted within the range of sediment Pb
concentration measured in the environment.
Saltwater Invertebrate Physiological Stress: Suggestive
Changes in antioxidant activity with Pb exposure are
reported in some saltwater invertebrates. The 2013 Pb ISA
included some evidence of invertebrate antioxidant
responses in bivalves and crustaceans at Pb concentrations
that are detected in the marine environment. Additional
evidence from environmental monitoring studies that
compared biomarker responses between reference and
contaminated sites indicated a correlation between the
amount of Pb with changes in antioxidant enzyme activity.
Studies published since the 2013 Pb ISA in saltwater
invertebrates, primarily mollusks, continue to show
perturbations to biomarkers of oxidative stress with
Pb exposure, albeit at concentrations of Pb higher
than typically countered in the environment
(Appendix 11.4.4.3.1).
Terrestrial Vertebrate Physiological Stress: Likely to Be Causal
Markers of oxidative damage are observed in terrestrial
mammals in response to Pb exposure.
The evidence since the 2013 Pb ISA
(Appendix 11.2.4.4.1) continues to support a likely to
be causal relationship between Pb exposure and
response in biomarkers of physiological stress. Most
new studies are in birds.
Freshwater Vertebrate Physiological Stress: Likely to Be Causal
Markers of oxidative damage are observed in fish and
amphibians in laboratory studies. Across freshwater
vertebrates, there are differences in the induction of
antioxidant enzymes with Pb exposure that appear to be
species-dependent.
Sublethal biomarker perturbations are associated with
Pb exposure in freshwater vertebrates
(Appendix 11.3.4.4.1.1). Few studies were identified
that reported physiological stress response at <10 |jg
Pb/L or concurrently assessed response at organism-
level endpoints.
Saltwater Vertebrate Physiological Stress: Inadequate
Insufficient evidence to assess causality
Insufficient evidence to assess causality
AQCD = Air Quality Criteria Document; ISA = Integrated Science Assessment; Pb = lead.
IS.8.4.2 Hematological Effects
As reported in the 2013 Pb ISA, inhibition of ALAD enzyme activity, an important rate-limiting
enzyme needed for heme production, is a recognized biomarker of Pb exposure (U.S. EPA. 2013a). The
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causality determinations for Pb effects on hematological endpoints in terrestrial, freshwater, and saltwater
organisms are unchanged from the 2013 Pb ISA (Table IS-16). Previous studies have indicated
considerable species differences in ALAD activity in response to Pb. At the time of the 2013 Pb ISA
evidence was sufficient to conclude that there is a causal relationship between Pb exposures and
hematological effects in terrestrial vertebrates and inadequate to assess causality between Pb exposures
and hematological effects in terrestrial invertebrates. Since the 2013 Pb ISA, additional species of birds,
amphibians and mammals have been shown to experience decreased ALAD activity following exposure
to Pb further supporting the existing causal relationship. For freshwater vertebrates, the evidence
evaluated in the 2013 Pb ISA and Pb AQCDs was sufficient to conclude that there is a causal relationship
between Pb exposures and hematological effects. Hematological effects of Pb on fish reported in the 2013
Pb ISA and AQCDs include a decrease in RBCs and inhibition of ALAD with elevated Pb exposure
under various test conditions. Inhibition of ALAD is also reported in environmental assessments of metal-
impacted habitats. In the 2013 Pb ISA it was determined that a causal relationship is likely to exist
between Pb exposures and hematological effects in freshwater invertebrates. Limited evidence from
saltwater invertebrates was suggestive of a causal relationship between Pb exposures and hematological
effects while evidence for saltwater vertebrates was insufficient to assess causality. Evidence for
hematological effects in saltwater invertebrates in previous AQCDs and the 2013 Pb ISA were primarily
from field monitoring studies of marine bivalves that used ALAD as a biomarker for Pb exposure and
correlated ALAD inhibition to increased Pb tissue content. Few new studies were identified since the
2013 Pb ISA that quantified ALAD response in terrestrial invertebrates or aquatic invertebrates or
vertebrates; hence causality relationships for hematological effects of Pb are unchanged.
Table IS-16 Summary of evidence for effects of Pb on hematological endpoints
in terrestrial and aquatic biota
Evidence from the 2013 Pb ISA Evidence from the 2024 Pb ISA
Terrestrial Invertebrate Hematological Effects: Inadequate
Insufficient evidence to assess causality Insufficient evidence to assess causality
Freshwater Invertebrate Hematological Effects: Likely to Be Causal
In metal-impacted habitats, ALAD is a recognized biomarker
of Pb exposure. Laboratory studies with freshwater
invertebrates have indicated considerable species
differences in ALAD activity in response to Pb. Field studies
in freshwater bivalves report a correlation between Pb and
ALAD activity.
No recent studies quantifying ALAD activity in
freshwater invertebrates at environmentally relevant
concentration of Pb were identified for inclusion in this
ISA.
Saltwater Invertebrate Hematological Effects: Suggestive
Field studies in saltwater bivalves report a correlation Few additional studies have reported inhibition of
between Pb and ALAD activity. ALAD activity in Pb-exposed saltwater invertebrates
and the concentrations at which enzyme activity is
affected appear to be much higher than Pb typically
encountered in seawater (Appendix 11.4.4.3.1).
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Evidence from the 2013 Pb ISA Evidence from the 2024 Pb ISA
Terrestrial Vertebrate Hematological Effects: Causal
In the 1986 Pb AQCD, decreases in RBC ALAD activity
were documented in birds and mammals near a smelter
(U.S. EPA. 1986b). Additional evidence for effects of Pb
blood parameters and their applicability as biomarkers of Pb
exposure in terrestrial birds and mammals were presented in
the 2005 Ecological Soil Screening Levels for Lead (U.S.
EPA. 2005). the 2006 Pb AQCD (U.S. EPA. 2006). and the
2013 Pb ISA. Field studies include evidence for elevated
BLLs correlated with decreased ALAD activity in songbirds
and owls living in historical mining areas.
Freshwater Vertebrate Hematological Effects: Causal
In the 1986 Pb AQCD, hematological effects of Pb exposure
in fish included decrease in RBCs and inhibition of ALAD
(U.S. EPA. 1986b). In fish, Pb effects on blood chemistry
have been documented with Pb concentrations ranging from
100 to 10,000 |jg Pb/L in studies cited in the 2006 Pb AQCD
(U.S. EPA. 2006).
Saltwater Vertebrate Hematological Effects: Inadequate
Insufficient evidence to assess causality Insufficient evidence to assess causality
AQCD = Air Quality Criteria Document; ALAD = 6-aminolevulinic acid dehydratase; BLL = blood lead level; ISA = Integrated
Science Assessment; RBC = red blood cell; Pb = lead.
IS.8.4.3 Neurobehavioral Effects
Organism-level endpoints include effects on behavior linked to Pb neurotoxicity. The causality
determinations for neurobehavioral effects of Pb in terrestrial, freshwater, and saltwater organisms remain
unchanged from the 2013 Pb ISA. Evidence for causality determinations for neurobehavioral endpoints
are summarized in Table IS-17. The 2013 Pb ISA concluded that the neurobehavioral effects of Pb
exposure on terrestrial and freshwater invertebrates are likely causal. In terrestrial invertebrates, the 2013
Pb ISA (U.S. EPA. 2013a) reported evidence of neurobehavioral aberrations such as impaired locomotion
in nematode Caenorhctbditis elegans that persisted over several generations while limited studies in
freshwater invertebrates provided evidence of decreased ability to escape or avoid predation in worms and
snails. Additional evidence since the 2013 Pb ISA in support of the likely to be causal relationship
between Pb exposure and neurobehavioral effects in terrestrial invertebrates include studies quantifying
alterations in foraging and feeding behavior in bees (Appendix 11.2.4.3.2). A few new studies including
effects on locomotion in amphipods and bivalves, and alternation of neurotransmitter activity in bivalves
and gastropods further support the 2013 finding of a likely to be causal relationship between Pb exposure
and neurobehavioral endpoints in freshwater invertebrates (Appendix 11.3.4.3.2). Evidence is inadequate
to assess causality between Pb exposure and neurobehavioral endpoints in saltwater invertebrates.
In the 2013 Pb ISA, the evidence was sufficient to conclude that the relationship between Pb
exposure and neurobehavioral effects in terrestrial and freshwater vertebrates is likely to be causal. Diet
New evidence (Appendix 11.2.4.4.1) continues to
support a causal relationship between Pb exposure
and hematological effects in terrestrial vertebrates
with most new evidence in birds. ALAD inhibition
correlated with increased blood Pb concentrations in
waterfowl, passerines, and scavengers as well as
livestock and toads.
Few studies were identified since the 2013 Pb ISA
that quantify ALAD response at concentrations
considered for this ISA. (Appendix 11.3.4.4.1).
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and injection studies in gull chicks and in lizards, designed to obtain Pb blood levels comparable to
organisms exposed in the wild, resulted in a variety of abnormal behaviors. For aquatic vertebrates,
evidence in prior AQCDs included behavioral impairment of a conditioned response in goldfish (U.S.
EPA. 1977) and several studies in which Pb was shown to affect predator-prey interactions in fathead
minnows (U.S. EPA. 2013a. 2006). Since the 2013 Pb ISA, there are additional studies on
neurobehavioral response particularly in zebrafish (Appendix 11.3.4.4.1.2). which are used as an animal
model for human health outcomes associated with Pb exposure. Endpoints assessed in zebrafish assays,
such as decreased locomotor activity and altered social interactions used as surrogates for autistic
behaviors in humans, can affect organism fitness in natural environments. Furthermore, some of these
studies link changes in gene expression, neurotransmitter levels or other molecular and cellular responses
to the observed behavioral outcomes. These new studies continue to support the likely to be causal
relationship between Pb exposure and effects on neurobehavior in aquatic vertebrates.
Table IS-17 Summary of evidence for effects of Pb on neurobehavioral
endpoints in terrestrial and aquatic biota
Evidence from the 2013 Pb ISA Evidence from the 2024 Pb ISA
Terrestrial Invertebrate Neurobehavioral Effects: Likely to Be Causal
A few studies reported altered feeding rates in snails while
others reported no effects. In limited studies available on
nematodes, there is evidence that Pb may affect the ability
to escape or avoid predation. Additional evidence of
changes in the morphology of GABA motor neurons was
also found in nematodes (C. elegans).
Nematode studies reported food preference, food
finding ability, and feeding activity were altered by Pb
exposure. New evidence in additional taxa include
findings that Pb negatively affects foraging and
feeding behavior as well as cognitive flexibility in bees
(Appendix 11.2.4.3.2).
Freshwater Invertebrate Neurobehavioral Effects: Likely to Be Causal
In the 2006 Pb AQCD, several studies were reviewed in
which Pb was shown to affect predator-prey interactions. In
limited studies available on worms and snails, there is
evidence that Pb may affect the ability to escape or avoid
predation.
A few studies further support the finding of a likely to
be causal relationship between Pb exposure and
neurobehavioral endpoints (Appendix 11.3.4.3.2).
These endpoints include effects on locomotion in
amphipods and alteration of neurotransmitter activity
and foot movement in a freshwater bivalve.
Saltwater Invertebrate Neurobehavioral Effects: Inadequate
Insufficient evidence to assess causality Insufficient evidence to assess causality
Terrestrial Vertebrate Neurobehavioral Effects: Likely to Be Causal
Limited behavioral studies in gull chicks experimentally
exposed to Pb reported abnormal behaviors such as
decreased walking, learning deficits, erratic behavioral
thermoregulation, and food begging that could make them
more vulnerable in the wild (Burger and Gochfeld. 2005).
Lizards exposed to Pb through diet exhibited abnormal
coloration and posturing behaviors. These results also
cohere with findings in laboratory animals that show that Pb
induces changes in learning and memory.
A few additional studies in birds since the 2013 Pb
ISA reported a relationship between Pb exposure and
neurobehavior or reported no effects
(Appendix 11.2.4.4.2). In one study in mockingbirds,
higher BLLs were correlated with increased levels of
aggressive behavior (McClelland etal.. 2019).
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Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Freshwater Vertebrate Neurobehavioral Effects: Likely to Be Causal
In the 2006 Pb AQCD, exposure to Pb was shown to affect
brain receptors in fish and may alter avoidance behaviors
and predator-prey interactions. Studies cited in the 2013 Pb
ISA included those that provided additional evidence for Pb
effects on behaviors that may impact predator avoidance
(swimming) and prey capture ability.
Several studies with larval zebrafish (Danio rerio)
bolster the finding of a likely to be causal relationship
from the 2013 Pb ISA. Some effects on behavioral
endpoints such as locomotion and social interactions
were reported at <20 |jg Pb/L (Appendix 11.3.4.4.1.2).
Saltwater Vertebrate Neurobehavioral Effects: Inadequate
Insufficient evidence to assess causality Insufficient evidence to assess causality
AQCD = Air Quality Criteria Document; BLL = blood lead level; GABA = gamma-aminobutyric-acid; Pb = lead.
IS.8.4.4 Survival
Survival may have a direct impact on population size and can lead to effects at the community
and ecosystem level of biological organization. Survival is commonly assessed in laboratory bioassays
and reported as a toxicological dose descriptor (e.g., 50% lethal concentration [LC50], LC20) to facilitate
comparison of effects across species and test conditions. In the 2013 Pb ISA the evidence was inadequate
to conclude that there is a causal relationship between Pb exposure and survival in terrestrial, freshwater,
or saltwater plants and this continues to be the case (Table IS-18). For invertebrates, the causality
determinations for survival remain unchanged from the 2013 Pb ISA (Table IS-18). At that time, the
evidence was sufficient to conclude that there is a causal relationship between Pb exposures and survival
in terrestrial and freshwater invertebrates and inadequate for saltwater invertebrates.
Terrestrial invertebrates typically tolerate high concentrations of Pb relative to concentrations
found in most uncontaminated soils. For freshwater invertebrates, key studies in amphipods reported in
the 2006 Pb AQCD and 2013 Pb ISA indicate a response to Pb at <10 |ig Pb/L under some water
conditions. Several studies since the 2013 Pb ISA provide further characterization for known effects on
survival in a few sensitive species of freshwater invertebrates, notably gastropods and amphipods, at
<15 (ig Pb/L in chronic exposures in which the concentration of Pb was analytically verified
(Appendix 11.3.5).
Evidence is sufficient to conclude that there is likely to be a causal relationship between Pb
exposure and survival in terrestrial vertebrates, with most of the direct evidence coming from studies of
waterfowl and birds of prey conducted throughout the last 50 years. For freshwater vertebrates, studies in
fish provided the basis for causal relationship for survival in the 2013 Pb ISA. Additional fish bioassays
conducted in varying water chemistry conditions report effects on survival at Pb concentrations similar to
those reported in the 2013 Pb ISA further supporting the causal relationship between Pb exposure and
survival in freshwater vertebrates (Table IS-18). Several additional studies have considered the effects of
Pb on native fish species including white sturgeon (Acipenser trcmsmontamis), and westslope cutthroat
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trout (Oncorhynchns clctrkii lewisi) (Appendix 11.3.5). Other recent studies on freshwater vertebrates
have further characterized the response to Pb under varying water conditions.
In the 2013 Pb ISA and previous assessments, the evidence for Pb effects on survival of saltwater
vertebrates was inadequate. New evidence (Appendix 11.4.5) is limited to laboratory-based bioassays in a
few fish species, toxicity data for other saltwater vertebrates remains lacking. Several recent chronic
bioassays conducted with early lifestages of three saltwater fish species report NOEC in the range of 11-
14 |ig Pb/L (Appendix 11. Table 11-7). Furthermore, Pb in these bioassays was analytically verified. In
the larval fish Topsmelt (Atherinops affinis), an LC50 = 15.1 |ig Pb/L; NOEC <13.8 |ig Pb/L was observed
at a salinity of 14 ppt (Reynolds et aL 2018). Calculated chronic values for additional saltwater fish
species that are consistent with the range reported above include grey mullet (Mugil cephahis) fingerling
survival and Tiger perch (Terctpon jarbna) fingerling survival (Hariharan et al.. 2016). Given these new
chronic studies in saltwater fish, the causality determination for this endpoint has changed since the 2013
Pb ISA and the evidence is suggestive of, but not sufficient to infer, a causal relationship between Pb
exposure and saltwater vertebrate survival (Table IS-18).
Table IS-18 Summary of evidence for effects of Pb on survival of terrestrial and
aquatic biota
Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Terrestrial Plant Survival: Inadequate
Insufficient evidence to assess causality
Insufficient evidence to assess causality
Freshwater Plant Survival: Inadequate
Insufficient evidence to assess causality
Insufficient evidence to assess causality
Saltwater Plant Survival: Inadequate
Insufficient evidence to assess causality
Insufficient evidence to assess causality
Terrestrial Invertebrate Survival: Causal
Survival of soil-associated invertebrates is adversely
affected by Pb exposure, albeit at Pb concentrations
associated with contaminated sites. In nematodes, the 2006
Pb AQCD reported LC50 values varying from 10 to 1,550 mg
Pb/kg dry weight dependent upon soil OM content and soil
pH (U.S. EPA. 2006). In earthworms, 14 and 28-d LC50
values typically fell in the range of 2,400 to 5,800 mg Pb/kg
depending upon the species tested. More recent evidence
has been consistent with these values and showed
concentration-dependent decreases in survival in
collembolans and earthworms under various experimental
conditions.
Evidence continues to support a causal relationship
between Pb exposure and invertebrate mortality,
although most reported effects occurred at
concentrations that greatly exceed environmental
concentrations. Additional bioassays include studies
in garden snails and earthworms
(Appendix 11.2.4.3.2).
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Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Freshwater Invertebrate Survival: Causal
Most of the evidence for Pb effects on survival in freshwater
invertebrates is from sensitive species of gastropods,
amphipods, cladocerans, and rotifers. In the 2006 Pb
AQCD, Pb impacted the survival of some aquatic
invertebrates at <20 |jg/L dependent upon water quality
variables (i.e., DOC, hardness, pH). Evidence in the 2013
Pb ISA also showed effects on survival in a few additional
freshwater invertebrates at <20 |jg Pb/L. Toxicity testing with
amphipods under various water conditions indicate these
organisms are sensitive to Pb at <10 |jg Pb/L.
Several studies provide further characterization for
known effects on survival in a few sensitive species of
freshwater invertebrates at <20 |jg Pb/L. In the
gastropod L. stagnalis, survival was significantly
decreased at 8.4 |jg Pb/L after 21-d exposure and
decreased survival was observed up to the end of a
56-d full lifecycle assessment (Munlev et al.. 2013). In
a chronic 42-d bioassay with the amphipod H. azteca,
survival was similar under two different experimental
diets conducted concurrently (LC20 = 15 [jg Pb/L and
LC20 =13 |jg Pb/L) (Besser et al.. 2016), and the
results supported the previous findings of effects in
amphipods in the low |jg/L range (Appendix 11.3.5
and Table 11-5).
Saltwater Invertebrate Survival: Inadequate
Limited evidence suggests that effects on survival are not
observed in most saltwater invertebrates unless Pb
concentrations greatly exceed those typically detected in
seawater.
Terrestrial Vertebrate Survival: Likely to Be Causal
Evidence continues to show that effects on survival
are typically not observed in bioassays unless Pb
concentrations greatly exceed those typically
detected in seawater.
In terrestrial avian and mammalian species, toxicity is
observed in laboratory studies over a wide range of doses
(<1 to >1,000 mg Pb/kg body weight per day) as reviewed
for the development of ecological soil screening levels (U.S.
EPA. 2005). and subsequently reported in the 2006 Pb
AQCD (U.S. EPA. 2006). The no-observed-adverse-effect
level for survival ranged from 3.5 to 3,200 mg Pb/kg per day.
No new studies were available within the scope of
this ISA reporting effects of Pb exposure on the
survival of terrestrial vertebrates.
Freshwater Vertebrate Survival: Causal
There is considerable historic information on Pb toxicity to
freshwater fish. Early observations from highly impacted
mining areas where Pb and other metals were present
indicated local extinction offish from streams (U.S. EPA.
1977). Several studies in the 2013 Pb ISA reported effects
at <100 |jg/Pb L in juvenile lifestages of a few fish species.
In the 2013 Pb ISA, 96-hr LC50 values in fathead minnow
tested in natural waters across the United States were as
low as 41 |jg Pb/L (Esbauqh etal.. 2011).
Additional fish bioassays conducted in varying water
chemistry conditions report effects on survival at Pb
concentrations similar to those in the 2013 Pb ISA
(Appendix 11.3.5 and Table 11-5). For larval
zebrafish (D. rerio), 96-hr LC50 values varied with
water hardness; in soft water LC50 = 52.9 |jg Pb/L
and in hard water LCso=>590 |jg Pb/L (Alsop and
Wood. 2011). In 96-hr acute tests conducted with two
lifestages of white sturgeon (Acipenser
transmontanus), the lowest 96-hr LCsowas 177 |jg
Pb/L in 8-d post-hatch larvae (Vardv et al.. 2014).
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Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Saltwater Vertebrate Survival: Suggestive (Inadequate in the 2013 Pb ISA)
Insufficient evidence to assess causality Additional evidence since the 2013 Pb ISA includes
chronic bioassays with analytically verified
concentrations of Pb conducted with early lifestages
in three saltwater fish species that report NOECs in
the range of 11-14 |jg Pb/L (Appendix 11.4.5 and
Table 11-7). In the larval fish Topsmelt (Atherinops
affinis), survival was impacted to a greater extent at
lower salinity (LCso = 15.1 |jg Pb/L; NOEC <13.8 |jg
Pb/L) than higher salinity (LCso = 79.8 |jg Pb/L;
NOEC = 45.5 |jg Pb/L) (Reynolds et al.. 2018).
Calculated chronic values for additional saltwater fish
species include a NOEC = 14 |jg Pb/L and
LOEC = 29 |jg Pb/L for grey mullet (Mugil cephalus)
fingerling survival and a NOEC = 11 |jg Pb/L and
LOEC = 22 |jg Pb/L for Tiger perch (Terapon jarbua)
fingerling survival following 30 d exposure to Pb
(Hariharan etal., 2016).
AQCD = Air Quality Criteria Document; d = day(s); DOC = dissolved organic carbon; hr = hour(s); ISA = Integrated Science
Assessment; LC50 = 50% lethal concentration; LOEC = lowest-observed-effect concentration, NOEC = no-observed-effect
concentration; OM = organic matter; Pb = lead.
IS.8.4.5 Growth
Alterations in the growth of an organism can impact population, community, and ecosystem-level
variables. In plants, the 2013 Pb ISA concluded that the relationship between Pb exposure and reduced
growth is causal in terrestrial plants and likely to be causal in freshwater aquatic plants. New evidence
continues to support these findings (Table IS-19). Evidence was inadequate for growth endpoints for
saltwater plants and algae in 2013 and this continues to be the case. There is evidence over several
decades of research that Pb inhibits photosynthesis and respiration in terrestrial plants, both of which
reduce growth (U.S. EPA. 2013a. 2006. 1977). Effects reported in plants are typically observed in
laboratory or greenhouse settings with high exposure concentrations or in field studies near stationary
sources and heavily contaminated sites, but studies that include multiple concentrations of Pb show
increased response with increasing concentration. In the 2006 Pb AQCD, half maximal effect
concentration (EC50) values for growth inhibition in various freshwater algal and aquatic plant species
were between approximately 1,000 and >100,000 |ig Pb/L and were mostly based on nominal
concentration data (U.S. EPA. 2006). An important advancement since the 2013 Pb ISA is the availability
of bioassay data for algal growth rate in several freshwater species based on measured Pb concentration
instead of nominal concentration, which strengthens confidence in the findings for the concentrations
assessed (Appendix 11.3.5). In conclusion, most primary producers experience EC50 values for growth at
concentrations that greatly exceed Pb concentrations typically found in U.S. soils and surface waters.
The 2013 Pb ISA concluded that the relationship between Pb exposure and decreased growth in
freshwater invertebrates is causal, and likely to be causal in terrestrial invertebrates. Building upon the
evidence for growth effects reported in the draft Ambient Aquatic Life Water Quality Criteria for Lead
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(U.S. EPA. 2008) and the 2006 Pb AQCD (U.S. EPA. 2006). studies reviewed in the 2013 Pb ISA
reported some effects at <10 (ig Pb/L for growth endpoints in aquatic invertebrates (U.S. EPA. 2013a).
The growth of the freshwater snail L. stagnctlis was identified as one of the most sensitive organisms and
endpoints for Pb toxicity. Since then, additional studies have supported previous findings of Pb effects on
the growth of this species at <10 (ig Pb/L lYCremazv et al.. 2018; Munlev et al.. 2013; Brix et al.. 2012;
Esbaugh et al.. 2012); Appendix 11. Table 11-5], The evidence remains inadequate to infer a causality
relationship for Pb exposure and reduced growth in saltwater invertebrates, and terrestrial and aquatic
vertebrates.
Table IS-19 Summary of evidence for growth effects of Pb in terrestrial and
aquatic biota
Evidence from the 2013 Pb ISA Evidence from the 2024 Pb ISA
Terrestrial Plant Growth: Causal
Effects of Pb on plant growth are typically observed in
laboratory studies with high exposure concentrations or in
field studies near stationary sources. In terrestrial plants,
there is evidence over several decades of research that Pb
inhibits photosynthesis and respiration, all of which can
reduce the growth of the plant (U.S. EPA. 2006. 1986a.
1977). The 2006 Pb AQCD relied principally on evidence
assembled in the Ecological Soil Screening Levels for Lead
document (U.S. EPA. 2005), which concluded that growth
(biomass) was the most sensitive and ecologically relevant
endpoint for plants. In the 2013 Pb ISA, there was some
evidence for exposure-dependent decreases in the biomass
of some plant species grown in Pb-amended or Pb-
contaminated soil.
Recent studies have continued to demonstrate
growth effects, albeit at concentrations that greatly
exceed Pb measured in soils. Growth endpoints
include decreases in photosynthetic performance,
damage to chlorophyll, increased antioxidant activity
in response to Pb stress, as well as genotoxic effects
of Pb. Studies of the effects of Pb on terrestrial plants
published since the last ISA continue to support the
previous known findings of declines in plant growth
under controlled exposures of Pb
(Appendix 11.2.4.2).
Freshwater Plant Growth: Likely to Be Causal
There is a large body of evidence to support growth effects
in plants at higher Pb concentrations. As reported in the
2013 Pb ISA and earlier AQCDs, there are documented
effects on growth in algae and aquatic plants in laboratory
studies. Most primary producers experience ECso values for
growth in the range of 1,000 to 100,000 |jg Pb/L,
concentrations that greatly exceed Pb concentrations
typically found in U.S. surface waters.
Saltwater Plant Growth: Inadequate
Saltwater species are historically underrepresented in
toxicity testing. In studies reviewed in the 2013 Pb ISA,
marine algae exhibited a range of sensitivity to Pb with a 72-
hr ECso reported for Chaetorceros sp. of 105 |jg Pb/L. Other
tested species were considerably less sensitive (72-hr
EC50 = 740 |jg Pb/L or higher).
Additional studies in algae and macrophytes continue
to support a likely to be causal relationship
(Appendix 11.3.4.1). A few new studies assessed the
sensitivity of freshwater algal growth to Pb exposure
and found a significantly negative effect in certain
species. New information on Pb effects on common
reed (P. australis) shows significant decreases in total
biomass, photosynthesis, and rhizome growth as well
as alterations in growth form and propagation
strategy under Pb exposure.
Limited evidence for growth inhibition for marine algal
species published since the 2013 Pb ISA, including a
few longer-term studies, generally show effects at
concentrations that greatly exceed environmental
concentrations (Appendix 11.4.4.2).
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Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Terrestrial Invertebrate Growth: Likely to Be Causal
A few studies cited in the 1986 Pb AQCD, the 2006 Pb
AQCD, and the 2013 Pb ISA reported growth effects in
terrestrial invertebrates and that effects were more
pronounced in juvenile organisms, underscoring the
importance of lifestage to overall Pb susceptibility. Some
studies also showed concentration-dependent inhibition of
growth in earthworms raised in Pb-amended soil.
Recent evidence continues to show growth-rate
effects in organisms associated with soil and food Pb
contamination including earthworms, snails, and
nematodes, as well as new evidence for tobacco
cutworm and fruit flies (Appendix 11.2.4.3.2).
Freshwater Invertebrate Growth: Causal
Some studies in sensitive freshwater invertebrates reported
inhibition of growth at or below 20 |jg Pb/L. The lowest
reported LOEC for growth in the 2006 Pb AQCD (16 |jg
Pb/L) was in amphipods (H. azteca) (Besser et al.. 2005). In
the 2013 Pb ISA, there was evidence for growth inhibition in
one species of snail (L. stagnalis) at <4 |jg Pb/L (Grosell and
Brix, 2009; Grosell et al.. 2006). The lowest genus mean
chronic toxicity value for Pb was 10 |jg Pb/L in a freshwater
mussel (Wang et al.. 2010).
Additional studies support previous findings of Pb
effects on growth of the snail (L. stagnalis) at <10 |jg
Pb/L (Cremazy et al., 2018; Munley et al., 2013; Brix
et al., 2012; Esbauqh et al., 2012) and a few other
invertebrates at or near 25 |jg Pb/L (Appendix 11.3.5
and Table 11-5).
Saltwater Invertebrate Growth: Inadequate
Insufficient evidence to assess causality
Insufficient evidence to assess causality
Terrestrial Vertebrate Growth: Inadequate
In AQCDs, growth effects of Pb have been reported in birds
(changes in juvenile weight gain) at concentrations typically
higher than currently found in the environment away from
heavily exposed sites.
No new studies were available within the scope of
this ISA reporting growth effects in terrestrial
vertebrates from Pb exposure.
Freshwater Vertebrate Growth: Inadequate
Evidence for growth effects of Pb is limited to a few studies
in amphibians and fish. Reports of Pb-associated growth
effects in freshwater fish are inconsistent; some studies
have shown no effects. Growth effects of Pb were reported
in frogs at concentrations typically higher than currently
found in the environment.
A few additional fish studies assessed growth
endpoints, with some reporting no effect
(Appendix 11.3.4.4.1.2).
Saltwater Vertebrate Growth: Inadequate
Insufficient evidence to assess causality
Few studies were identified since the 2013 Pb ISA
that assessed growth in saltwater vertebrates.
AQCD = Air Quality Criteria Document; EC50 = half maximal effect concentration; ISA = Integrated Science Assessment;
LOEC = lowest observed effect concentration; Pb = lead.
IS.8.4.6 Reproduction
Evidence from invertebrate and vertebrate studies from Pb AQCDs, the 2013 Pb ISA and in this
review indicates that Pb is affecting reproductive performance in multiple species (Table IS-20). Various
endpoints have been measured in multiple taxa of terrestrial and aquatic organisms to assess the effect of
Pb on development, fecundity, and hormone homeostasis, and they have demonstrated the presence of
adverse effects. Reproductive effects are important when considering effects of Pb because impaired
fecundity at the organism level of biological organization can result in a decline in abundance and/or
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extirpation of populations, decreased taxa richness, and decreased relative or absolute abundance at the
community level (Suter et al.. 2004). The evidence is inadequate to conclude that there is a causal
relationship between Pb exposures and developmental and reproductive effects in either terrestrial or
aquatic plants. In the 2013 Pb ISA the evidence was sufficient at that time to conclude that there is a
causal relationship between Pb exposures and developmental and reproductive effects in terrestrial and
freshwater invertebrates. New evidence suggests that in earthworms, Pb exposure slows the time to
maturation and that in isopods, it delays onset of the breeding season and shortens its duration, and that it
influences mate selection in fruit flies (Appendix 11.2.4.3.2). For freshwater invertebrates, recent
evidence further supports previous observations of Pb effects on reproductive endpoints at low
concentrations in sensitive species of gastropods, cladocerans and rotifers, especially under chronic
exposure scenarios (Appendix 11.3.5 and see Table 11-5).
In the 2013 Pb ISA, evidence was suggestive of a causal relationship between Pb exposure and
reproductive and developmental effects in saltwater invertebrates based on endpoints including delay in
onset to reproduction in amphipods, impaired larval development and embryogenesis inhibition in
bivalves, and a decrease in fertilization rate of eggs in a marine polychaete (U.S. EPA. 2013a). Since the
2013 Pb ISA, the evidence base for Pb effects on reproductive and developmental endpoints in saltwater
invertebrates has expanded, primarily due to multiple new embryo-larval developmental assays in
mollusca and echinodermata (Appendix 11.4.5 and Table 11-7). Several of these acute exposure bioassays
analytically verify the concentration of Pb at which effects were observed (Markich. 2021; Romero-
Murillo et al.. 2018; Nadella et al.. 2013) and report effects at lower concentrations than reported in the
2013 Pb ISA. Considering coherence of reproductive and developmental effects of Pb across species,
observations in saltwater invertebrates are consistent with terrestrial and freshwater invertebrates (both
"causal" in the 2013 Pb ISA). As a result of the newly available evidence since the 2013 Pb ISA the
causality determination for this endpoint has changed and the evidence is sufficient to infer a likely to be
causal relationship between Pb exposure and reproductive and developmental effects in saltwater
invertebrates.
In the 2013 Pb ISA, the evidence was sufficient to conclude that there is a causal relationship
between Pb exposures and developmental and reproductive effects in terrestrial and freshwater
vertebrates, and this continues to be the case. For reproduction and development in freshwater vertebrates,
the weight of evidence for the causal relationship in the 2013 Pb ISA was primarily from studies with
fish. Previous Pb AQCDs have reported reproductive and developmental effects in fish, including brook
trout (Salvelinus fontinalis), rainbow trout (Oncorhvnchus mvkiss), and fathead minnow (Pimephales
promelas) (U.S. EPA. 2013a. 2006. 1977). Other supporting evidence for the causal determination in the
2013 Pb ISA for reproductive effects in aquatic vertebrates included alteration of steroid profiles and
additional reproductive variables, although most of the available studies were conducted using nominal
concentrations of Pb. New early lifestage fish studies, including several in zebrafish (D. rerio) in which
the concentration of Pb in exposure water was analytically verified (Appendix 11.3.4.4.1.2) add to the
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existing evidence for Pb effects on endocrine and developmental endpoints. These studies at analytically
verified concentration of Pb include several developmental studies in amphibians (Appendix 11.3.4.4.3).
Table IS-20 Summary of evidence for reproductive effects of Pb in terrestrial
and aquatic biota
Evidence from the 2013 Pb ISA Evidence from the 2024 Pb ISA
Terrestrial Plant Reproduction: Inadequate
Insufficient evidence to assess causality Insufficient evidence to assess causality
Freshwater Plant Reproduction: Inadequate
Insufficient evidence to assess causality Insufficient evidence to assess causality
Saltwater Plant Reproduction: Inadequate
Insufficient evidence to assess causality Insufficient evidence to assess causality
Terrestrial Invertebrate Reproduction: Causal
The 2006 Pb AQCD reported effects on reproduction in
collembolans and earthworms, with LOECs and NOECs
typically well above Pb soil concentrations observed away
from stationary sources of contamination. In the 2013 Pb
ISA, studies in a few taxa expanded the evidence for Pb
effects on developmental and reproductive endpoints for
invertebrates at concentrations that generally exceed Pb
levels in U.S. soils. Evidence of multigenerational toxicity
effects of Pb is also present in terrestrial invertebrates,
specifically springtails, mosquitoes, carabid beetles, and
nematodes in which decreased fecundity in the progeny of
Pb-exposed individuals was observed.
Freshwater Invertebrate Reproduction: Causal
Reproductive effects of Pb in freshwater aquatic
invertebrates are well characterized in previous Pb AQCDs
and the 2013 Pb ISA and have been observed at or near
current ambient concentrations in some species in
laboratory exposures. Results under controlled conditions
have consistently shown reproductive effects of Pb in
sensitive taxa, especially amphipods and cladocerans, at
concentrations near Pb quantified in freshwater
environments.
Studies published since the 2013 Pb ISA continue to
support a causal relationship between Pb exposure
and invertebrate reproductive endpoints including
time to maturation and brood size
(Appendix 11.2.4.3.2). In addition to new studies in
earthworms and nematodes, additional new taxa
demonstrating reproductive effects associated with
Pb exposure include isopods and fruit flies. Several
multigenerational fruit fly studies together report that
Pb exposure influences female mate selection,
oviposition site, and tolerance to Pb contamination is
greater in populations with a history of Pb exposure.
Recent evidence (Appendix 11.3.5) further
characterizes Pb effects on reproductive endpoints at
low (<10 |jg Pb/L) concentrations in sensitive species
of gastropods, cladocerans, and rotifers
(Appendix 11. Table 11-5), especially under chronic
exposure scenarios.
Saltwater Invertebrate Reproduction: Likely to Be Causal (Suggestive of, but Not Sufficient to Infer
Causality in the 2013 Pb ISA)
For saltwater invertebrates, there is limited evidence for
effects on reproduction and early development. Reported
effects included a delay in the onset to reproduction in
amphipods (Rinaenarv et al.. 2007). impaired larval
development (Wang et al.. 2009) and embryogenesis
inhibition (Wang et al.. 2009: Beiras and Albentosa. 2004) in
bivalves and a decrease in the fertilization rate of eggs
(marine polycheate annelid) (Gopalakrishnan et al.. 2008).
These effects were observed for Pb concentrations higher
than typically detected in marine environments.
Multiple new embryo-larval developmental assays in
mollusca (mussels, oysters) and echinodermata (sea
urchin) (Appendix 11.4.5 and Table 11-7) have
expanded the evidence for reproductive effects since
the 2013 Pb ISA. Several of these acute exposure
bioassays analytically verified the concentration of Pb
at which effects were observed (Markich, 2021;
Romero-Murillo et al.. 2018; Nadella et al.. 2013) and
reported effects at lower effect concentrations than
those reported in the 2013 Pb ISA. For example, the
48-hr EC10 was 9-10 |jg Pb/L in two mussel species,
and 72-hr EC10 was 19 |jg Pb/L in sea urchin
Strongylocentrotus purpuratus (Nadella et al.. 2013).
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Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Terrestrial Vertebrate Reproduction and Development: Causal
Effects reported in the 2006 Pb ISA included declines in
clutch size, number of young hatched, number of young
fledged, decreased fertility, and decreased eggshell
thickness observed in birds near areas of Pb contamination
and in birds with elevated Pb tissue concentration
regardless of location (U.S. EPA. 2006). In the 2013 Pb ISA,
studies in a few taxa expand the evidence for Pb effects on
mammalian developmental and reproductive endpoints.
Recent studies, although limited, continue to support
a causal relationship between Pb exposure and
reproductive effects in terrestrial vertebrates. New
studies provide additional evidence of Pb exposure
causing decreased lifetime breeding success, lower
nestling weight at birth, decreased eggshell
thickness, and decreases in egg yolk antioxidant
levels in birds (Appendix 11.2.3.4.2).
Freshwater Vertebrate Reproduction and Development: Causal
The weight of evidence for reproductive and developmental
effects in freshwater vertebrates is from fish. Pb AQCDs
have reported developmental effects in a few fish species at
or near 120 pg Pb/L (U.S. EPA. 1977) (U.S. EPA. 1986b)
and reported effects on other reproductive endpoints
including decreased spermatocyte development (U.S. EPA.
2006). Reproductive effects in fish are influenced by water
chemistry.
Several studies in fish further support previous
findings of Pb effects on reproductive endpoints in
freshwater vertebrates (Appendix 11.3.4.4.1.2). A few
of these studies report effects at lower concentrations
than the 2013 Pb ISA or prior AQCDs. Specifically,
hatching success rates in zebrafish embryos were
reduced at 4.5, 9.6 and 18.6 pg Pb/L aqueous
exposure; (Zhao et al.. 2020). Endocrine disruption
(significant reduction in thyroid hormones
triiodothyronine (T3) and thyroxine (T4)) was
observed in zebrafish larvae following exposure to 30
pg Pb/L (Zhu etal.. 2014).
Saltwater Vertebrate Reproduction and Development: Inadequate
Insufficient evidence to assess causality
Insufficient evidence to assess causality
AQCD = Air Quality Criteria Document; EC10 = effect concentration at 10% inhibition; hr = hour(s); ISA = Integrated Science
Assessment; LOEC = lowest observed effect concentration; NOEC = no-observed-effect concentration; Pb = lead.
IS.8.4.7 Community and Ecosystem Effects
Endpoints relevant to assessing effects of Pb on communities and ecosystems include the
alteration of species richness, species composition, and biodiversity. Uptake of Pb into aquatic and
terrestrial organisms and subsequent effects on mortality, growth, development, and reproduction at the
organism level can cascade to effects on populations and communities and lead to ecosystem-level
consequences. Although the evidence is strong for the effects ofPb on growth, reproduction, and survival
in certain species in experimental settings, considerable uncertainties exist in generalizing effects
observed under experimental conditions and at a smaller scale to predicted effects at the community and
ecosystem levels of biological organization. In many cases, it is difficult to characterize the nature and
magnitude of ecosystem-level effects and to quantify relationships between environmental concentrations
of Pb and ecosystem response due to the presence of multiple stressors, variability in field conditions, and
differences in Pb bioavailability. In addition, although the presence of Pb is associated with shifts in
biological communities, this metal rarely occurs as a sole contaminant in natural systems, making the
contribution of Pb to the observed effects difficult to ascertain.
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a likely to be causal
relationship between Pb exposure and terrestrial and freshwater-community and ecosystem effects, and
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recent evidence continues to support this determination (Appendix sections 11.2.6. 11.3.6. 11.2.4.1. and
11.3.4.1 and Table IS-21). In terrestrial habitats, communities and ecosystems exposed to elevated Pb
concentration, typically from proximity to historically active metal extracting and processing point
sources, have been shown to suffer from decreased species diversity and changes in species composition.
These changes affect microbial, floral, and faunal communities. Since the 2013 Pb ISA, effects of Pb
exposure on the interactions between trees and their pests, between herbaceous plants and insects, and
plants, worms, and soils invertebrates have been added to the evidence (Appendix 11.2.6). Reductions in
species abundance, richness, or diversity associated with the presence of Pb in freshwater habitats are
reported in the literature, usually in heavily contaminated sites where Pb (and other metal) concentrations
are higher than typically observed environmental concentrations. Most evidence is from sediment-
associated microbial and macroinvertebrate communities. Since the 2013 Pb ISA (U.S. EPA. 2013a).
several experimental and observational studies have examined the relationship between Pb concentration
in the sediment and effects on freshwater microbes (Appendix 11.3.4.1). Several of these studies report
negative relationships between sediment Pb concentration and microbial abundance or community
structure, while some report no relationship or positive associations. Observational and experimental
studies published since the 2013 Pb ISA continue to show negative associations between sediment and/or
porewater Pb concentration and macroinvertebrate communities (Appendix 11.3.6).
For saltwater ecosystems, evidence was inadequate in the 2013 Pb ISA to assess causality
between Pb exposures and community and ecosystem effects. Reduced species abundance and
biodiversity of protozoan and meiofauna communities were observed in laboratory microcosm studies
with marine water and marine sediments reviewed in the 2006 Pb AQCD as summarized in Table AX7
2.5.2 (U.S. EPA. 2006). In the 2013 Pb ISA, there were a few additional studies including effects on
community structure and nematode diversity (U.S. EPA. 2013a). Since that time, there are new
experimental and observational studies (Table IS-21) examining the relationship between Pb in sediment,
and microbial abundance and/or diversity (Appendix 11.4.4.1). and Pb associations with saltwater
foraminiferal communities (Appendix 11.4.6). Several of the benthic foraminifera studies report effects
on community richness, diversity, and abundance. In other studies with foraminifera, there were changes
in abundance of certain taxa associated with Pb, but not diversity metrics. Considering the new evidence,
Pb quantified in sediment is a factor explaining variation in microbial diversity and foraminiferal species
distributions and abundance in a variety of distinct geographic locations. In these studies, Pb is often
correlated with other heavy metals. In addition to the available studies assessing Pb effects on saltwater
communities, primarily foraminifera, the effects of Pb on reproduction and survival of early lifestages in
sensitive saltwater invertebrates, especially when considered cumulatively, could impact populations, and
community and ecosystem structure and function. Population, community, or ecosystem-level studies are
typically conducted at sites that have been contaminated or adversely affected by multiple stressors
(several chemicals alone or combined with physical or biological stressors), which increase the
uncertainty of attributing observed effects to Pb. Therefore, additional evidence available since the 2013
Pb ISA indicates the evidence is suggestive of, but not sufficient to infer, a causal relationship between
Pb exposure and saltwater community and ecosystem effects.
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Table IS-21 Summary of evidence for community and ecosystem effects of Pb
Evidence from the 2013 Pb ISA
Evidence from the 2024 Pb ISA
Terrestrial Community and Ecosystem Effects: Likely to Be Causal
Independent effects of Pb are difficult to interpret because of
the presence of other stressors, including metals. The 1986
Pb AQCD (U.S. EPA. 1986a) reported shifts toward Pb-
tolerant communities at 500 to 1,000 mg Pb/kg soil. In the
2006 Pb AQCD (U.S. EPA. 2006), decreased species
diversity and changes in community composition were
observed in ecosystems surrounding former smelters. In the
2013 Pb ISA, there was additional evidence for Pb effects
on soil microbial communities.
Experimental studies have shown that trophic transfer
of Pb can affect species interactions, nematode
community composition, and bacterial abundance
and/or activity (Appendix 11.2.4.1). Additional
observational studies reported negative or null
relationships between soil Pb concentration and
microbial and invertebrate abundance and diversity
(Appendix sections 11.2.4.1 and 11.2.6).
Freshwater-Community and Ecosystem Effects: Likely to Be Causal
Effects of Pb are difficult to interpret because of the
presence of other stressors, including metals. Most evidence
of community and ecosystem-level effects is from near Pb
sources, usually mining effluents. In the 2013 Pb ISA
evidence for Pb effects on sediment-associated and
freshwater aquatic plant communities added to the existing
body of evidence of effects of Pb at higher levels of
biological organization.
Several studies reported negative correlations
between sediment Pb concentration and invertebrate
community composition or ecosystem processes
(Appendix 11.3.6). Additionally, observational and
experimental studies have reported negative
relationships between sediment and/or porewater Pb
concentration and microbial abundance and/or
community structure, while some reported no
relationship or positive associations
(Appendix 11.3.4.1)
Saltwater Community and Ecosystem Effects: Suggestive of, but Not Sufficient to Infer, a Causal
Relationship (Inadequate in the 2013 Pb ISA)
No studies on community and ecosystem-level effects of Pb
in marine systems were reviewed in the 1977 Pb AQCD
(U.S. EPA. 1977). or the 1986 Pb AQCD (U.S. EPA. 1986a).
Observations from field studies reviewed in the 2006 Pb
AQCD (U.S. EPA. 2006) included findings of a negative
correlation between Pb and species richness and diversity
indices of macroinvertebrates associated with estuary
sediments. Evidence for community and ecosystem-level
effects of Pb in saltwater ecosystems in the 2013 Pb ISA
included a few laboratory microcosm studies as well as
observations from field-collected sediments, biofilm, and
plants in which changes in community structure were
observed; however, evidence was inadequate to make a
causality determination at the time.
Additional studies since the 2013 Pb ISA provide
sufficient evidence for effects on saltwater
communities and ecosystems to be suggestive of a
causal relationship. Several studies report reductions
in foraminiferal and/or meiofaunal community
richness, diversity, and/or abundance associated with
higher concentrations of Pb in sediment and water,
while others found positive or null correlations
(Appendix 11.4.6). In addition, several experimental
and observational studies reported negative
relationships between sediment and/or saltwater Pb
concentrations and microbial abundance and/or
diversity, while other studies found no relationship
(Appendix 11.4.4.1).
AQCD = Air Quality Criteria Document; ISA = Integrated Science Assessment; Pb = lead.
IS.9 Policy-Relevant Issues
In the process of evaluating the current state of the science with respect to the effect of Pb
exposure on health and welfare, studies that conducted analyses that address some of the key policy-
relevant questions of this review were identified, as detailed in Volume 2 of the Pb IRP (U.S. EPA.
2022a). such as:
• To what extent has new information altered scientific conclusions regarding the relationships
between Pb in ambient air and Pb in children's blood?
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• To what extent does the newly available evidence alter our understanding of the C-R relationships
between Pb in children's blood and reduced IQ?
• To what extent is there new scientific evidence available to improve our understanding of the
health effects associated with various time periods of Pb exposures at various stages of life?
• Has new information altered our understanding of human populations that are particularly
sensitive to the current low environmental Pb exposures, including air-related exposures?
• Does the newly available evidence identify new endpoints or indicate new exposure levels at
which ecological systems or receptors are expected to experience effects?
The following sections summarize the evidence that can inform consideration of these policy-
relevant questions, specifically: air Pb-to-blood Pb relationships (Section IS.9.1); C-R relationship
between BLLs and IQ (Section IS.9.2); the level, timing, duration, and frequency of Pb exposure
contributing to observed health effects (Section IS.9.3), and populations potentially at increased risk of a
PM-related health effect (Section IS.9.4). A summary of recent evidence related to at-risk populations is
provided in Section IS.7.4.
IS.9.1 Air Pb-to-Blood Pb Relationships
Multivariate regression models, commonly used in epidemiology, provide estimates of the
variability in BLLs (or other biomarkers) explained by various exposure pathways (e.g., air Pb
concentration, surface-dust Pb concentration). Models can provide estimates of the rate of change of
blood or bone Pb concentration in response to an incremental change in exposure level (i.e., slope factor).
Within the literature and U.S. EPA documents, the relationship between air Pb and blood Pb is commonly
characterized in terms of a "slope factor" or "air-to-blood ratio." An air-to-blood ratio of 1:5 indicates that
for every 1 (ig/m3 of air Pb, there is a 5 (ig/dL increase in blood Pb. Synonymously, this is characterized
by a slope factor of 5 (ig/dL per |ig/nr\ Air Pb-blood Pb relationships in children, described in
Appendix 2.5.1. are summarized below.
The 1986 Pb AQCD (U.S. EPA. 1986a) described epidemiologic studies of relationships between
air Pb and blood Pb. Drawing from the studies examined, the aggregate blood Pb-air Pb slope factor
(when considering both air Pb and Pb in other media derived from air Pb) was estimated to be
approximately double the slope estimated from the contribution due to inhaled air alone (U.S. EPA.
1986a). Much of the pertinent earlier literature (e.g., prior to 1984, when air Pb was dominated by the use
of leaded gasoline in on-road motor vehicles) on children's BLLs was summarized by Brunekreef (1984)
and found that the blood Pb versus air Pb slope was smaller at high blood and air levels. Most studies
have empirically modeled the air Pb-to-blood Pb relationship using nonlinear regression (i.e., log-log),
which itself gives an increasing slope with decreasing air Pb concentration.
In the 2008 final rule for the Pb NAAQS (73 FR 66964), the Administrator's decision on the
revised level for the new primary standard was informed by an evidence-based framework for considering
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air-related IQ loss for children living near Pb sources. U.S. EPA, recognizing uncertainty and variability
in the air-to-blood relationships, interpreted the evidence as providing support for a range of estimates
inclusive of 1:5 at the lower end and 1:10 at the upper end, with the ratio of 1:7 identified as a central
estimate within the range supported by the evidence at the time (73 FR 67001-67002, 67005).
At the time of the 2013 Pb ISA (U.S. EPA. 2013a). there was uncertainty, due to the limited
evidence, in projecting the magnitude of the air Pb-blood Pb relationship to ambient air Pb below
0.2 |ig/m \ There are studies since the 2013 Pb ISA that evaluate the air Pb-to-blood Pb relationship that
are more reflective of current conditions with central tendency air Pb concentration (PbA) between 0.004
and 0.04 |ig/nr\ As was the case for older data in the 1986 Pb AQCD (U.S. EPA. 1986a). newer data also
show slope factors increasing with decreasing air Pb (Figure IS-4). Although saturable gastrointestinal
absorption and saturation of Pb binding to RBC occur at relatively high rates of Pb intake leading to
blood Pb concentration (PbB) of 20-30 (ig/dL (see Section 2.2). the nonlinear relationship between PbA
and PbB cannot be explained by a biokinetic mechanism.
100 x
Zahran etal. (2013)
. #2 years
1 year# '
3 years
^yearslo*
4 years O ,
6 years ^
7 years#
Meng etal. (2014), 1-5 years, PM10
years
Meng et al. (2014), 6-11 years, PM10
^ Rjchrnond-Bryant et al. (2013, 2014)
1999-2008 1988-1994
8-10 years# , 6-11 vrs# -1 -5 years
1-5 yrs# 6-11 years
Bierkens et al. (2011), <6 years
Meng et al. (2014), 1 -5 yrs, TSP
Ranft et al. (2008), 6-11 years, 3.0 |jg/dL#
Brunekreff (1984), PbB<20*g/dL
^^Schwartz and Pitcher (1989), 0-5 years
^ Hayes et al. (1994), 0.5-5 years
Hilts et al. (2003), 0.5-5 years
• Brunekreff (1984), all children
#Tripathi et al. (2001), 6-10 years
Meng et al. (2014), 6-11 years, TSP
Ranft et al. (2008), 6-11 years, 1.5 |jg/dL
+
0.001
0.01
0.1
Air Pb (ng/m3)
10
Source: Figure based on Richmond-Bryant et al. (2014) with data from Table 2-13 (Appendix 2).
Figure IS-4 Slope factors for blood Pb as a function of air Pb.
IS-103
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In general, longitudinal studies conducted after phasing out leaded gasoline would best inform the
current relationship of blood Pb to air Pb. Ideally, such studies would compare two populations for which
air Pb concentrations differ while all other Pb sources are unchanged. In a nearly ideal study, Hilts (2003)
reported the change in blood Pb from 1996 to 2001 for children under 5 years old associated with the
emission reduction from a local smelter in Trail, BC, Canada. However, even in this study, the reduction
in exposure from pathways other than air cannot be ruled out because of the "comprehensive education
and case management programs/' An advancement in analyses of the blood Pb-air Pb association came
from leveraging the U.S. EPA Air Quality System (AQS) with NHANES surveys. The blood Pb-air Pb
associations across different NHANES periods should reflect the change in this association for the U.S.
population over time (Richmond-Bryant et al.. 2014; Richmond-Bryant et al.. 2013) because each
NHANES cycle is a representative sample of the U.S. population. However, merging blood Pb results
from multiple NHANES periods with the U.S. EPA AQS could introduce exposure measurement errors
as well as uncertainties in terms of population representativeness and availability of covariates. Each
single study presented in Table 2-13 (Appendix 2.5.1) deviates from the ideal design in one or more
aspects. Collectively, all of these studies contribute to our understanding of how air Pb impacts blood Pb.
IS.9.2 Concentration-Response Relationships for Human Health Effects
In assessing the relationship between Pb exposure and human health effects, an important
consideration is the shape of the C-R relationship across the full range of Pb biomarker levels and
whether there is a threshold concentration below which there is no evidence of an effect. As described
elsewhere in the document (Appendix 2.3). the interpretation of the epidemiologic study findings depends
on the exposure history of the study populations and the choice of the biomarker in the context of what is
known about that exposure history. Many of the adult populations examined in older and more recent
epidemiologic studies are likely to have had higher past than recent Pb exposure. Given their longer
exposure histories, there is uncertainty regarding the frequency, duration, timing, and level of exposure
contributing to the blood Pb or bone Pb levels measured in adult and adolescent populations. Specifically,
higher past exposures may bias C-R estimates based on later childhood, adolescent, and adult BLLs.
Therefore, this section summarizes evidence relevant to thresholds and C-R relationships in studies of
childhood Pb exposure. A summary of previous evidence regarding C-R relationships for exposure
biomarkers in adolescents and adults is presented in Section 1.9.3 of the 2013 Pb ISA (U.S. EPA. 2013a).
and a summary of recent evidence can be found within the health effects appendices.
With each previous assessment (U.S. EPA. 2013a. 2006). the epidemiologic and toxicological
evidence demonstrated that progressively lower BLLs or Pb exposures are associated with cognitive
deficits in children. The 2006 Pb AQCD found that cognitive effects in children were observed in
association with BLLs of 10 (ig/dL and lower, while the evidence assessed in the 2013 Pb ISA found that
an association between BLLs and cognitive effects in children was substantiated to occur in populations
with mean BLLs between 2 and 8 (ig/dL. The conclusion of the 2013 Pb ISA was based on studies that
IS-104
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examined early childhood BLLs (i.e., age <3 years), considered peak BLLs in their analysis (i.e., peak
<10 (ig/dL), or examined concurrent BLLs in young children (i.e., age 4 years). A recent study of
Canadian preschool children from generally middle- to upper-middle SES families with low Pb exposure
(mean concurrent BLL = 0.70 (ig/dL) did not find an association between concurrent Pb exposure and IQ
at age 3-4 years (Desrochers-Couture et al.. 2018). Although some other recent studies report
associations between Pb exposure and cognitive effects in children with mean BLLs <2 (ig/dL, recent
studies generally include somewhat older children, or employ modeling strategies designed to answer
relatively narrow research questions (e.g., the effect of joint exposure to Pb and other metals, or the effect
of concurrent Pb exposure independent from prenatal exposure) and consequently do not have the
attributes of the studies on which the conclusion of the 2013 Pb ISA was based (i.e., early childhood
BLLs, consideration of peak BLLs, or concurrent BLLs in young children). Furthermore, studies that
might extend the evidence related to exposure-response relationships (i.e., recent studies that reflect the
lower early childhood Pb exposures now more common in the United States with an adequate range of Pb
exposure [i.e., studies of subjects with BLLs <1 to 2 (ig/dL measured during relevant time periods]) are
limited. Overall, the recently available studies were not designed, and may not have the sensitivity, to
detect (Cooper et al.. 2016) the effect or hazard at these very low BLLs, nor do they provide evidence of a
threshold for the effect across the range of BLLs examined.
Epidemiologic studies evaluated in the 2013 Pb ISA provided evidence of a larger decrement in
cognitive function per unit increase in blood Pb among children with lower mean BLLs compared with
children with higher mean BLLs. Key evidence was provided by an international pooled analysis of seven
prospective cohort studies (Lanphear et al.. 2019. 2005). as well as studies that examined prenatal or early
childhood BLLs or considered peak BLLs in school-aged children or concurrent BLLs in young children
(i.e., 2 years old). Recent studies that evaluate the shape of the C-R function for the relationship between
Pb exposure and cognitive effects in children are limited in number, but continue to support the
conclusions from the 2013 Pb ISA. In particular, a re-analysis of the pooled data set of Lanphear et al.
(2005). Crump et al. (2013) corroborated the finding that there was evidence of a nonlinear C-R function
over the range of the BLLs evaluated (e.g., 2.5-33.2 (ig/dL, as 5th to 95th percentile concurrent BLLs) -
i.e., a larger incremental effect of Pb exposure on IQ at lower blood Pb concentrations as indicated by a
log-linear C-R function. Lanphear et al. (2005) also fit linear functions over stratified BLL ranges (e.g., <
7.5 (ig/dL and >7.5 (ig/dL) that similarly indicated statistically significantly larger Pb-associated cognitive
function decrements across the lower range compared to the higher range. Individual studies also support
this finding, showing greater decrements in cognitive function per unit increase in BLL among children in
lower strata of BLLs compared with children in higher strata of BLLs [Figure 4-15, and Table 4-16 of
U.S. EPA (2013a)l. Notably, uncertainty in the shape of the C-R relationship increases at lower BLLs due
to a smaller number of observations. Previous assessments also noted attenuation of C-R relationships at
higher exposure or dose levels in the occupational literature. Reasons proposed to explain the attenuation
include greater exposure measurement error and saturation of biological mechanisms at higher levels, as
well depletion of the pool of susceptible individuals at higher exposure levels (Stavner et al.. 2003).
Possible explanations specific to nonlinear relationships observed in studies of Pb exposure in children
IS-105
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include a lower incremental effect of Pb due to covarying risk factors such as low SES, poor caregiving
environment, and higher exposure to other environmental factors (Schwartz. 1994). differential activity of
mechanisms at different exposure levels, and confounding by omitted variables or misspecified variables
(U.S. EPA. 2013a). Review of the evidence did not reveal a consistent set of covarying risk factors to
explain the differences in the blood Pb-IQ C-R relationship across high and low Pb exposure groups
observed in epidemiologic studies. Additionally, although evidence indicates a larger incremental effect
of Pb exposure on IQ at lower BLLs, consistent findings of higher mean IQ at lower BLLs indicates that
the absolute magnitude of the effect of Pb exposure on cognitive function declines with decreasing BLL.
IS.9.3 Lifestages and Timing of Pb Exposure Contributing to Observed
Nervous System Effects
As discussed in Appendix 2.3.5. blood Pb may reflect both recent as well as past exposures
because Pb is both taken up by and released from the bone. The relative proportion of BLLs resulting
from recent versus past exposure is uncertain in the absence of specific information about the pattern of
exposure contributing to observed BLLs, which is generally not ascertainable in epidemiologic studies.
This uncertainty is greater in adults and older children, than in young children who do not have lengthy
exposure histories. As a result, stronger conclusions can be reached regarding the timing of exposures that
result in health effects in children. Several lines of evidence, which are summarized below, inform the
interpretation of epidemiologic studies of young children with regard to the patterns of exposure that
contribute to observed health effects.
The collective body of epidemiologic evidence reviewed in the 2013 Pb ISA did not provide
strong evidence to identify an individual critical lifestage or timing of Pb exposure with regard to
neurodevelopmental effects in children (U.S. EPA. 2013a). Specifically, epidemiologic studies reviewed
in the 2013 Pb ISA consistently showed that BLLs measured during various lifestages and time periods
(i.e., prenatal, early childhood, childhood average, and concurrent with the outcome) were associated with
nervous system effects in children. Recent studies generally support this conclusion, though several
studies indicate that increases in postnatal (earlier childhood, lifetime average, concurrent) BLLs were
associated with larger cognitive function decrements in children ages 4-10 years than increases in
prenatal BLLs. These results suggest that per unit increase, postnatal Pb exposures that are reflected in
concurrent or cumulative BLLs or tooth Pb levels may have a larger magnitude of effect on cognitive
function decrements as children age. Notably, however, exposure metrics that are based on blood Pb
measurements at different ages in childhood are typically highly correlated. Consequently, the relative
importance of various exposure metrics, which are measured during different lifestages and time periods,
is difficult to discern in epidemiologic studies. Evidence in rodents and monkeys, however, indicates that
Pb exposures during multiple lifestages and time periods, including prenatal only, prenatal plus
lactational, postnatal only, and lifetime are observed to induce impairments in learning (Rice. 1992; Rice
and Gilbert. 1990). Additionally, recent prospective epidemiologic studies observed associations between
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childhood BLLs and decrements in IQ during late adolescence (18-19 years) and mid-adulthood (38-
45 years). These findings provide insight into the persistence of Pb-associated cognitive function
decrements and are consistent with the understanding that the nervous system continues to develop
(i.e., synaptogenesis and synaptic pruning remains active) throughout childhood and into adolescence.
IS.9.4 Ecological Effects and Corresponding Pb Concentrations
Pb that is released into air, soil, or water is then cycled through any, or all, of these media before
reaching an ecological receptor. When a plant, invertebrate, or vertebrate is exposed to Pb, the proportion
of observed effects attributable to Pb from atmospheric sources is difficult to quantitatively assess
because of a lack of information not only on deposition but also on bioavailability, as affected by specific
characteristics of the receiving ecosystem, and on the kinetics of Pb distribution in long-term exposure
scenarios. Although long-term trends in declining anthropogenic emissions of Pb are detected in some
non-air media and biota, the connection between air concentration and ecosystem exposure continues to
be poorly characterized for this metal, and measurements of the contribution of atmospheric Pb to specific
sites is generally unavailable. Current evidence indicates that Pb is bioaccumulated in biota, however, the
sources of Pb in biota have only been identified in a few studies, and the relative contribution of Pb from
each source is usually not known.
No new endpoints were identified for Pb effects in terrestrial, freshwater, or saltwater biota since
the 2013 Pb ISA. However, a few effects were reported at lower concentration than for the 2013 Pb ISA,
primarily in chronic laboratory-based bioassays for endpoints that were already established as causal in
the 2013 Pb ISA. The level at which Pb elicits a specific effect continues to be difficult to establish in
terrestrial and aquatic systems. There are large differences in species sensitivity to Pb, and many
environmental variables (e.g., pH, OM) determine the bioavailability and toxicity of Pb.
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EPA/600/R-23/375
APDA Environmental Protection Januaiy 2024
M m Agency www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 1: Lead Source to Concentration
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
-------
CONTENTS
DOCUMENT GUIDE 1-iii
LIST OF TABLES 1-v
LIST OF FIGURES 1-vi
ACRONYMS AND ABBREVIATIONS 1-vii
APPENDIX 1 LEAD SOURCE TO CONCENTRATION 1-1
1.1 Introduction 1-1
1.2 Sources of Atmospheric Pb 1-2
1.2.1. Aviation Gasoline and Aircraft Emissions 1-7
1.2.2. Industrial Sources 1-9
1.2.3. Fuel Combustion 1-11
1.2.4. Fires 1-12
1.2.5. Traffic and Roads 1-15
1.2.6. Volcanoes 1-17
1.2.7. Legacy Sources 1-18
1.2.8. Other Sources 1-20
1.3 Fate and Transport of Pb Emitted into Air 1-21
1.3.1. Fate and Transport in Air 1-21
1.3.2. Fate and Transport in Soil 1-24
1.3.3. Fate and Transport in Water and Sediments 1-30
1.3.4. Fate and Transport in Urban Media 1-48
1.4 Monitoring of Pb in Ambient Air 1-53
1.4.1. Network Monitoring 1-53
1.4.2. Federal Reference Methods 1-56
1.4.3. Sampling Considerations 1-58
1.4.4. Recent Advances in Sampling and Analysis 1-59
1.5 Ambient Air Pb Concentration Trends 1-61
1.5.1. National Scale Ambient Air Concentrations and Long-Term Trends 1-62
1.5.2. Seasonal and Diurnal Trends 1-70
1.5.3. Particle Size Characteristics 1-72
1.5.4. Background Concentrations 1-74
1.6 Summary and Conclusions 1-74
1.7 References 1-75
1-iv
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LIST OF TABLES
Table 1-1 Annual lead emissions (tons) from aircraft operating modes 1-9
Table 1-2 Parameters related to fires and associated Pb measurements discussed herein 1-15
Table 1-3 Distribution of regulatory Pb-total suspended particle concentrations in |jg/m3 for 2020-
2022 1-64
Table 1-4 Distribution of Pb concentrations for various types of measurements and monitoring site
locations in |jg/m3 for 2020-2022 1-64
Table 1-5 Seasonal variations in Pb concentration in ambient air 1-70
1-v
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LIST OF FIGURES
Figure 1-1 U.S. Pb emissions by sector. 1-3
Figure 1-2 U.S. county-level Pb emissions density estimates in Ibs/year/mile2 1-4
Figure 1-3 U.S. anthropogenic Pb emissions trend 1990-2020. 1-5
Figure 1-4 The biogeochemical cycle of tetramethyl/tetraethyl Pb. 1-22
Figure 1-5 Ambient air Pb and air soil concentrations and median splines in |jg/m3 from Detroit,
Michigan. 1-51
Figure 1-6 Map of U.S. ambient air monitoring sites reporting regulatory Pb data to U.S. EPA during
the 2020-2022 period. 1-55
Figure 1-7 Map of U.S. ambient air monitoring sites reporting non-regulatory Pb data to U.S. EPA
during the 2020-2022 period. 1-56
Figure 1-8 Pb maximum rolling 3-month average in pg/m3 for the 2020-2022 period. 1-63
Figure 1-9 Site-level trends in maximum rolling 3-month average Pb concentrations for 2010-2022. 1-65
Figure 1-10 National trend in Pb annual maximum 3-month average concentrations in |jg/m3, 2010 to
2022. 1-66
Figure 1-11 Distribution of annual maximum 3-month concentrations measured at regulatory Pb
monitoring sites, 1980 to 2022. 1-67
1-vi
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ACRONYMS AND ABBREVIATIONS
As arsenic
AQCD Air Quality Criteria Document
AQS Air Quality System
Ca calcium
CEC cation exchange capacity
CFR Code of Federal Regulations
CSN Chemical Speciation Network
DI deionized
DO dissolved oxygen
DOC dissolved organic carbon
DOM dissolved organic matter
ED-XRF energy-dispersive X-ray fluorescence
spectrometry
EF enrichment factor
FAAS flame atomic absorption spectroscopy
Fe iron
FEM Federal Equivalent Method
FRM Federal Reference Method
FTC freeze thaw cycle
FIA humic acid
FlAP hazardous air pollutant
ICP-MS inductively coupled plasma mass
spectrometry
IMPROVE Interagency Monitoring of Protected
Visual Environments
K potassium
Me-L metal ligand
Mg magnesium
Mn manganese
NAAQS National Ambient Air Quality
Standards
NATTS National Air Toxics Trends Stations
NCore National Core multipollutant
monitoring network
NEI National Emissions Inventory
OM organic matter
OX oxide-bound
PM particulate matter
PTFE polytetrafluoroethylene
SLAMS state and local air monitoring stations
SPM suspended particulate matter
SS/CAR specifically sorbed/carbonate-bound
STR soil temperature regimes
Ti titanium
TRI Toxic Release Inventory
TSS total suspended solids
TSP total suspended particulate
UFP ultrafine particle
U.S EPA United States Environmental Protection
Agency
Zn zinc
1-vii
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APPENDIX 1 LEAD SOURCE TO CONCENTRATION
1.1 Introduction
This appendix characterizes the current state of atmospheric and environmental science relevant
to understanding lead (Pb) exposure and Pb-related health and ecological effects described in subsequent
appendices. It builds on previous research reviewed in the 2013 Integrated Science Assessment for Lead
(hereinafter referred to as the 2013 Pb ISA) (U.S. EPA. 2013) and previous Pb Air Quality Criteria
Documents (AQCDs) (U.S. EPA. 2006. 1986. 1977) and emphasizes relevant advances in sources and
emissions, fate and transport, sampling and analysis methods, and concentration trends. Because of the
large body of literature on the subject, this appendix focuses primarily on new studies from the United
States and Canada, with a few exceptions for highly relevant international publications. The scope is not
limited to airborne Pb from contemporary emission sources because non-atmospheric processes, as well
as legacy sources, are also relevant for understanding the effects of airborne Pb. For example, transport
and transformation processes in soil and water after deposition are also relevant. Therefore, current
research in other media is also included to promote understanding of airborne Pb in the context of non-
atmospheric sources and media.
In previous ISAs, an up-to-date review of air emissions, monitoring, and concentration trends has
been accomplished through a combination of analysis of United States Environmental Protection Agency
(U.S. EPA) ambient air monitoring network data and a synthesis of observations reported in the peer-
reviewed literature. Reference data such as total emissions, coverage of network monitors, average
concentrations, and concentration trends can become out of date before the document is published
because these data are so frequently updated. To facilitate provision of the most current emissions and
concentration data from the Pb monitoring network, a set of relevant maps and graphics that have been
routinely included in the atmospheric appendix or chapter in previous ISAs are now drawn from a
separate document "Overview of Lead (Pb) Air Quality in the United States." Many of the figures and
tables in this appendix are from the 2022 Overview (U.S. EPA. 2022a). These will be updated annually
and made available at https://www.epa.gov/air-qualitv-analvsis/lead-naaqs-review-analvses-and-data-sets.
This appendix complements the Overview by providing a literature-based synthesis of recent research on
Pb sources, fate and transport, measurement, and concentration trends. Section 1.2 provides an overview
of sources and emissions of Pb in ambient air and other environmental media. Section 1.3 gives
descriptions of the fate and transport of Pb in air, soil, and aqueous media. Section 1.4 describes advances
in Pb measurement methods, and Section 1.5 describes ambient air Pb concentrations, including spatial
and temporal variability on national and local scales and the size distributions of Pb-bearing particulate
matter (PM).
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1.2 Sources of Atmospheric Pb
Figure 1-1 summarizes the major sources of U.S. Pb emissions based on the 2020 National
Emissions Inventory (NEI) (U.S. EPA. 2023a). which is divided into mobile and stationary sources. Total
estimated national Pb emissions for 2020 were 621 tons, with 69% accounted for by aircraft, 18% from
industrial processes, 9% from stationary fuel combustion, and 3% from fires. All other sources combined
were estimated to account for about 2% of total U.S. Pb emissions. The sum of emissions from all
industrial sources of 218,000 lbs agreed well with total national stack emissions from the Toxic Release
Inventory (TRI) of 210,000 lbs for 2020 (U.S. EPA. 2023d). Not shown in Figure 1-1 are emissions from
resuspended soil. Resuspension of Pb from historical sources has not been quantified in this way on a
national scale but has been demonstrated to be a dominant source in urban areas where it has been studied
and estimates of Pb resuspension near roads and industrial sources have been reported in a number of
studies described in previous assessments (U.S. EPA. 2013. 2006) and in Section 1.3.4.
Figure 1-2 shows the geographic distribution of Pb emissions from the 2020 NEI. High county
emissions often correspond to counties that also experienced the highest ambient concentrations described
in Section 1.5.1. Total Pb emissions have steadily decreased for decades, largely due to the elimination of
leaded gasoline use in automobiles before 1996 and in later years because of reductions in emissions from
metals processing sources (U.S. EPA. 2013. 2006). As shown in Figure 1-3, from 1990 to 2020, there has
been a steep decline in total U.S. Pb emissions from about 5 kton/year to less than 1 kton/year and a
replacement of industrial sources with non-road mobile sources as the dominant category of emissions,
which reflects the prominence of aircraft using leaded aviation fuel as the largest emissions source (U.S.
EPA. 2023c)). Up-to-date graphics of total U.S. Pb emissions estimates by source, geographic distribution
ofPb emissions estimates, and the 30-year total U.S. emissions estimates trends are available in
"Overview of Lead (Pb) Air Quality in the United States" (U.S. EPA. 2022a). and updated annually at
https://www.epa.gov/air-qualitv-analvsis/lead-naaqs-review-analvses-and-data-sets.
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Pb Emissions (621 Tons/year)
Source: (U.S. EPA. 2023c).
Figure 1-1 U.S. Pb emissions by sector.
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Lead Emissions Density in lbs/year/miA2 (# Counties)
~ 0-0.29 (2,398) ~ 0.3-0.99 (559) ~ 1-2.99 (204) ~ 3-9.99 (50) ¦ 10-31 (10)
Source: (U.S EPA. 2023c).
Figure 1-2 U.S. county-level Pb emissions density estimates in
Ibs/year/mile2.
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Inventory Year
Source: (U.S. EPA. 2023c).
Figure 1-3 U.S. anthropogenic Pb emissions trend 1990-2020.
The NEI is a national compilation of emissions information provided by state, local, and tribal air
agencies, as well as source sector emission estimates developed by the U.S. EPA. Uncertainties are not
reported here because the inventory database does not include sector emissions uncertainty estimates. The
intent is to create the most complete inventory for use in air quality modeling, national rule assessments,
international reporting, and other reports. It focuses largely on anthropogenic sources, with information
about natural sources where available. The NEI program develops data sets, blends data from multiple
sources, and performs quality assurance steps that further enhance and augment the compiled data. The
accuracy of individual emission estimates may vary from facility to facility or county to county, and for
some sources, data may be incomplete or lacking. For example, there are no soil resuspension data in the
NEI. Sources of error may include the measurements used for developing source-specific emissions
factors, biases related to modeled activity levels, or sources missing from the inventory. While
uncertainties are difficult to predict, the NEI undergoes continuous improvement by the U.S. EPA with
the assistance of state, local, and tribal agencies by their reporting emissions information for facilities,
other stationary sources, and mobile sources. Each 3-year cycle of NEI development incorporates
improvements based on lessons learned from the previous cycles, and estimation procedures for emissions
sectors typically evolve over time in response to identified deficiencies as the data are used. As a result, in
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spite of inexact and potentially unknown uncertainties, the NEI largely meets the needs for general
emissions assessments and national trends reporting. Quality assurance procedures and acceptance criteria
for the NEI are detailed in the NEI technical support document (U.S. EPA. 2023a).
While emissions inventory data are essential for understanding emissions, there are potential
limitations and uncertainties. Uncertainty estimates for major emission sources or comparison of
uncertainties between different types of sources were not found in the recent literature, and this was also
the case in previous Pb air quality criteria documents and the 2013 Pb ISA. Harris et al. (2006) pointed
out it is generally not clear how facilities calculate emissions reported to the NEI or TRI and that they do
not attempt to quantify uncertainties. A comparison across several inventories covering the same area
found the inventories sometimes did not include all emission sources, contained data that were not
current, and reported emissions that varied considerably within the same year, leading to a
recommendation that emissions data would benefit from data sharing, greater uncertainty analysis, and
standardization of emissions estimation methods (Harris et al.. 2006).
Source attribution studies have been carried out to investigate the extent to which different
sources contribute to environmental concentrations either in remote areas to distinguish between major
source types or at or near industrial sites to distinguish between component industrial processes for a
given source. The basic approach was described in detail in the 2013 Pb ISA (U.S. EPA. 2013). Isotopes
of Pb are measured in source samples to obtain a library of source signatures that can then be compared to
ambient samples to estimate source contributions when an ambient sample is analyzed. As examples,
isotope ratios vary between gasoline, coal, smelter, and municipal waste emissions, and specific ratios can
ultimately be traced back to geologic conditions under which the Pb associated with each source was
formed. As described in the 2013 Pb ISA, isotope ratios were first applied to PM in the mid-1960s to
identify the impact of motor vehicle exhaust on urban Pb deposition. Several Pb isotope source attribution
studies were reviewed in the 2013 Pb ISA, including both general source attribution studies and studies
targeting quantification of specific source contributions, including coal and waste incineration. In more
recent years, there has been little research activity on ambient air, but a continued focus on quantifying
specific source contributions, especially smelting (Section 1.2.2), wildfires (Section 1.2.4), and legacy Pb
from leaded gasoline (Section 1.2.7). Recent source attribution studies using a Pb isotope ratio approach
have also been applied to Pb deposition (Section 1.3.1.2), soil (Section 1.3.2.1), and sediments
(Section 1.3.3.4).
For context, much of the Pb in the United States is neither emitted into air nor transported into air
from other media. Non-air Pb sources include plumbing (Santucci and Scully. 2020; Frank et al.. 2019;
USGS. 2018; Rosen et al.. 2017; Stillo and Macdonald Gibson. 2017; Hanna-Attisha et al.. 2016; Pieper
et al.. 2015; Brown et al.. 2011). mine waste (Duval et al.. 2020; Gutierrez et al.. 2020; Pavlowskv et al..
2017). and food (FDA. 2022; Martin-Domingo et al.. 2017; Ritchie and Gerstenberger. 2013; Gunev and
Zagurv. 2012). Even airborne Pb is only partly produced by contemporary Pb emissions into the
atmosphere. Contemporary sources are discussed in this section, including aviation gasoline and aircraft
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emissions (Section 1.2.1), industrial emissions (Section 1.2.2), fuel combustion from stationary sources
(Section 1.2.3), wildland fires (Section 1.2.4), traffic-related emissions (Section 1.2.5), and volcanoes
(Section 1.2.6). Substantial contributions to airborne Pb can also be attributed to historical sources of
airborne Pb (Section 1.2.7) and non-atmospheric Pb sources, the Pb from which can in some cases
become airborne through the processes of suspension and resuspension (Section 1.3.4). The resulting
airborne Pb concentrations observed in ambient air (Section 1.5) can thus potentially be the result of a
combination of contemporary atmospheric sources, resuspension of historical atmospheric sources and
non-atmospheric sources.
1.2.1. Aviation Gasoline and Aircraft Emissions
Leaded aviation gasoline, or avgas, used by piston engine aircraft, is the largest national source of
Pb emitted into the atmosphere identified by the NEI and is responsible for 69% of atmospheric Pb
emissions in the United States. Pb additives, usually in the form of tetraethyl Pb, prevent engine knocking
that could result in sudden engine failure (U.S. EPA. 2013). Most avgas is considered "100 Low Lead",
which contains 2.12 g Pb/gallon (ASTM. 2021; U.S. EPA. 2013). In 2017, 208M gallons of avgas were
consumed in the U.S. (FAA. 2020). leading to -470 tons of total Pb emissions (U.S. EPA. 2022a). In
2020, the COVID-19 pandemic caused a 60% reduction in air traffic (Sher et al.. 2021).
At the time of the last review, it was reported that Pb emissions from aircraft come in gaseous or
particulate forms (U.S. EPA. 2013). In aviation exhaust PM, Pb is largely composed of lead bromide
(PbBr2) crystals coated with hydrocarbons (Griffith. 2020; U.S. EPA. 2013). These particles are typically
under 100 nm in diameter, although they can form larger agglomerates (Turgut et al.. 2020). Pb particles
emitted from piston engine aircraft exhaust have been observed as small as 13 nm diameter, which are
significantly smaller than the mode of particles emitted from vehicle exhaust (35 nm) (Griffith. 2020). Pb
had the highest concentration of any element measured by inductively coupled plasma mass spectrometry
(ICP-MS) in PMio collected directly from aircraft engine exhaust ducts (median Pb value of
4.6 x 106 ng/m3) and was 40 times more concentrated than the next most abundant element (Na) (Turgut
et al.. 2020). Avgas constituents, including tetraethyl Pb, can evaporate into the headspace of storage and
fuel tanks or be exhausted from the engine in the gas phase (NASEM. 2021; U.S. EPA. 2013). Annual
evaporative emissions of Pb from refueling are estimated at 75 kg (NASEM. 2021).
Around a single airport at which leaded fuel is used, Pb in air is highest around the runways
(Rahim et al.. 2019; Turgut et al.. 2019; Feinberg et al.. 2016; Feinberg and Turner. 2013). The U.S. EPA
used an estimate of 7.34 g of Pb emissions during a single take-off and landing cycle to estimate airport
Pb emissions in the NEI (U.S. EPA. 2013). Touch-and-go operations are commonly practiced during pilot
training and account for up to 23% of flights, depending on the airport (U.S. EPA. 2020b). Touch-and-go
operations generally remain in air near the airport and can involve the aircraft circling overhead for hours,
potentially contributing Pb to the local environment near airports used for training (U.S. EPA. 2020b).
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Aircraft run-up, the series of checks performed by pilots immediately prior to take-off, contributed up to
82% of the 3-month average Pb concentration at one airport modeled by the U.S. EPA (U.S. EPA.
2020b). Aircraft engine run-up has been identified as one of the most important emission sources for
ground-level Pb concentrations and was estimated at one airport to burn approximately 15.3 g/second of
fuel and 50 g/second of fuel for a single- and multiple-engine aircraft, respectively (U.S. EPA. 2013; Carr
et al.. 2011). Median run-up times measured at one airport were 40 and 63 seconds for single- and
multiple-engine aircraft, respectively (U.S. EPA. 2020b). These times correspond to a three-month
average Pb concentration of 0.092 (ig/m3, though this measurement was only conducted for one three-
month period at one airport (U.S. EPA. 2020b). Model-extrapolation analyses for airports using leaded
avgas estimated 3-month average Pb concentration at the site of maximum impact for some airports with
high landing and take-off activity, to be <1-475 ng/m3 at one airport (U.S. EPA. 2020b) and 10-20 ng/m3
at another airport (Feinberg and Turner. 2013). Fuel consumption estimated at one airport during taxi
ranged from 1.6 g/second to 5.1 g/second for single- and multiple-engine aircraft, respectively (U.S. EPA.
2013; Carr et al.. 2011). The time an aircraft spends in taxi can have a significant influence on Pb
concentrations, as taxi time can vary greatly (Feinberg et al.. 2016). Taxiing was responsible for 12% of
total Pb emissions reported in one study, with about half of those emissions occurring when the aircraft
was idle and awaiting clearance for take-off (Feinberg et al.. 2016). These studies indicate runways to be
the primary hot spot for Pb emissions at airports that use avgas. Annual aircraft Pb emissions from each
flight phase are reported in Table 1-1.
There are 13,117 airports and over 5,000 heliports in operation in the United States (NASEM.
2021). Remote states, such as Montana and Alaska, rely heavily on air transportation. Alaska contains
nearly 10% of the total amount of airports in the United States (NASEM. 2021). Several studies have
observed lower Pb emissions when air traffic is lower (Rahim et al.. 2019; Zahran et al.. 2017). There
have also been observations of significantly decreased Pb concentrations near airports during periods of
precipitation compared to when it is dry (Rahim et al.. 2019).
Soil and air Pb concentrations decrease with distance from an airport. Model-extrapolated
airborne Pb concentrations decrease with increasing distance from runway, from -0.6 (.ig/rn3 at the
maximum impact site to 0.01 (.ig/rn3 500 m from the runway (U.S. EPA. 2020b). Soil samples collected
from an Oklahoma airport were analyzed for Pb, for which elevated soil Pb concentrations were generally
observed within 500 m from an airport (McCumber and Strevett. 2017). However, a few sites
demonstrated higher soil Pb concentrations more than 500 m from an airport, suggesting influence from
other sources such as industrial, historical, or non-air sources. This study also identified hot spots (10-
170 mg Pb/kg) near fueling centers, suggesting avgas as the primary source (McCumber and Strevett.
2017). Twenty-four-hour average air Pb concentrations ranging from 17-70.6 ng/m3 were reported and
remained above background levels (10 ng/m3) up to 900 m from a Santa Monica airport (Carr et al..
2011). illustrating the possible dispersion of Pb emissions from aircraft. At the same airport, PM2 5 Pb
concentrations dropped from 24 ng/m3 to ~6 ng/m3 after shortening the runway by 450 m (Hudda et al..
2022). This 75% reduction in airborne Pb concentration was attributed to a 50% decrease in aviation
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operations following the shrinking of airport size (Hudda et al.. 2022). Ambient air Pb was measured in a
U.S. EPA one-year monitoring study at 17 U.S. airports for a full year ending in December of 2013 (U.S.
EPA. 2015). Monitoring was required for a set of airports with estimated Pb emissions between 0.50 and
1.0 tons Pb per year that also met additional criteria including the dominant use of one runway and the
level of piston-engine aircraft activity. Airport Pb concentrations monitored depend on level of piston-
engine aircraft activity, the patterns of runway use, meteorology, and the placement of the monitor
relative to the run-up area, and other factors. Maximum 3-month average Pb concentrations ranged from
0.1 to 0.33 (ig/m3 and exceeded 0.15 (ig/m3 at 2 of the 17 airports (U.S. EPA. 2015). In general, blood
lead levels have been shown to increase with proximity to airports, volume of piston-engine aircraft
traffic, and avgas consumption at airports (Zahran et al.. 2023). These results are discussed further in
Section 2.4.1.
Unleaded gasoline for motor vehicles became available following government-sponsored
programs like the Federal Motor Vehicle Control program (U.S. EPA. 2022b). To reduce Pb emissions
from aviation, unleaded avgas has recently been developed for most piston engine aircraft (Swift Fuels.
2023; Bertorelli. 2021; AVweb. 2013).
Table 1-
1 Annual lead emissions (tons) from aircraft operating modes
Aircraft
type
Taxi-Out Run-Up
Take-off
Climb
Approach
Landing
Taxi-in Total
Multiple
engine
1.31 x 10"3 6.27 x 10"4
3.78 x 10"4
4.07 x 10"4
1.86 x 10"4
8.51 x 10"5
2.20 x 10"4 3.21 x 10"3
Single
engine
9.94 x10"3 4.37 x 10"3
3.74 x 10"3
4.43 x 10"3
3.69 x 10"3
8.00 x 10"4
2.72 x 10"3 2.97 x10"3
Source: (U.S
2]
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CM
O
(N
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LU
1.2.2. Industrial Sources
The 2013 Pb ISA summarized emissions inventory data and source apportionment results that
attributed substantial amounts of airborne Pb to the metals industry (U.S. EPA. 2013). Studies identifying
primary smelting mainly of other metals from various ores, secondary smelting—mainly of Pb batteries,
and steel manufacturing as important contributors to Pb emissions were reviewed. Observations from
several publications of downwind airborne Pb concentrations and Pb-PM content from the last remaining
primary Pb smelter in the United States were summarized, as well as emissions from smelters for other
metals that continue to operate. Other industrial emissions contributed less than the metals industry at the
time of the 2013 Pb ISA, and there were few publications and little discussion of their emissions and
contribution.
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According to the 2020 NEI, industrial sources now account for 18% of U.S. emissions, making
them the second largest category of sources after aviation fuel (U.S. EPA. 2023a). Roughly half of
industrial emissions are from metal industries, both ferrous and non-ferrous in approximately equal
amounts, and emissions sources include smelters, steel mills, foundries, and metal fabrication operations.
The other half of industrial emissions is not related to metals processing and includes industries such as
glass and cement manufacturing. Recent research on industrial Pb sources and emissions is largely limited
to a few high-profile areas. Since publication of the 2013 Pb ISA, there have been several studies
published on various aspects of Pb emissions from smelters, but there are few recent studies on other
industrial Pb emissions, whether metals related or otherwise.
The highest ambient air Pb concentrations in the United States in Figure 1-2 are observed near
metal industry sources. Historically, some large Pb smelters have been among the largest single sources
of U.S. Pb emissions. Together they dominated local Pb emissions and accounted for a large fraction of
national Pb emissions after the removal of Pb from gasoline (U.S. EPA. 2013). Recent studies have also
continued to identify specific smelters as major urban Pb sources (Wang et al.. 2011). Numerous previous
field studies have documented Pb emissions from smelters as well as elevated ambient air Pb
concentrations in the vicinity of primary smelters and soil Pb concentrations decreasing with distance
from smelters (Bowers et al.. 2014; U.S. EPA. 2013). Recent research generally confirms these earlier
observations by also showing that soil Pb concentrations decreased with distance from North American
smelters and that isotope ratios consistent with smelter emissions could be identified in soil some distance
from the smelter (Widorv et al.. 2018; Felix et al.. 2015).
One major focus of recent research has been Pb size distributions of smelter emissions. A
bimodal particle size distribution with maxima at 0.18 |im and 9.9 |im was consistently observed over
several years of sampling in the vicinity of a large copper (Cu) smelter in Hayden AZ (U.S. EPA. 2013;
Csavina et al.. 2011). Csavina et al. (2014) confirmed that airborne Pb followed a bimodal particle size
distribution in the vicinity of industrial operations that had both mining and smelting operations in both
Arizona and Australia and suggested that the finer particles (<1 |im) were produced from smelters and the
coarser particles were from windblown dust sources like mine tailings, crushing and grinding operations,
and regional or nearby urban sources. Pb isotope ratios were used to show that fine particles smaller than
1 (mi aerodynamic diameter in the vicinity of large smelters were mainly due to emissions from the
smelter, while the isotope signature of coarse particles near the smelter were more similar to PM from a
regional background aerosol or nearby urban environments (Felix et al.. 2015; Csavina et al.. 2014). For
similar mining operations in the absence of a smelter, only a coarse mode was observed (Csavina et al..
2014). Fugitive emissions of airborne dust studied using a Bayesian framework also reinforced that
substantial Pb emissions are associated with coarse PM. Pb associated with airborne dust from loading
and storage areas were estimated from time-dependent airborne Pb concentration measurements in
multiple locations in the vicinity of a Pb-zinc (Zn) smelter in Trail, British Columbia, and found to make
significant contributions to Pb emissions (Hosseini and Stockie. 2016). As substantiated by results from
multiple smelters, large smelting operations can be a large local source of airborne Pb in both fine and
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coarse PM, and Pb emissions from smelters can also have a broad area of impact because of their
concentration in the fine particle size range (Csavina et al.. 2014).
There have also been advances in describing the physical and chemical properties of Pb in
smelter emissions. Previous speciation data from smelter emissions reviewed by U.S. EPA (2006) and
Skeaff et al. (2011) are qualitative or semi-quantitative in nature (Skeaff et al.. 2011). Skeaff et al. (2011)
set as their objective the development of a quantitative chemical speciation of stack particulates from
three Cu smelters with amass balance as close to 100% as possible using X-ray diffraction, scanning
electron microscopy, and electron probe microanalysis. Acceptable mass balances were achieved, and Pb
accounted for 7.5% to 14% of PM by weight across the three smelters. Although insoluble PbSC>4 was
consistently the dominant form of Pb (Skeaff et al.. 2011). another study found that in the vicinity of a
smelter in Hayden AZ, the PM size range most enriched in Pb overlapped with the most hygroscopic PM
mode (Youn et al.. 2016; Sorooshian et al.. 2012).
1.2.3. Fuel Combustion
Fuel combustion contributes -55 tons of Pb/year to the atmosphere (9% of total emissions) with
the greatest contributions from coal and oil fired boilers (20 tons) and coal-fired electric power generation
(15 tons) (U.S. EPA. 2023a). Previous reports have provided extensive background on the role of Pb in
coal combustion. Pb is found in coal in varying amounts (5-35 mg Pb/kg) (U.S. EPA. 2013. 2006). Fly
ash, a byproduct of coal combustion, is composed primarily of silicon and oxygen (Zierold and Odoh.
2020; U.S. EPA. 2006). Pb in fly ash is enriched 2-10 times compared with that in parent coal (Zierold
and Odoh. 2020; Wang et al.. 2019) for concentrations in fly ash samples ranging from 25.3-308 mg
Pb/kg (Wang et al.. 2019) or 1.4-2120 mg Pb/kg depending on the source (Zierold and Odoh. 2020). In a
study performed in Colorado and the Appalachian Basin, 54% of Pb from parent coal was found in fly ash
particles at concentrations of 41.8 mg Pb/kg (Swanson et al.. 2013).
Pb emissions from coal-fired power plants have decreased by 36% since 1993 due to pollution
control measures and plant closures, though power plants can still dominate local Pb emissions (U.S.
EPA. 2020a; Zierold and Odoh. 2020; Gingerich et al.. 2019). In New Mexico, average Pb concentrations
in PM2 5 samples were 0.65 ng/m3 (range 0.20-1.04 ng/m3), with approximately 0.44 ng/m3 attributed to
the two nearby coal-fired power plants (Gonzalez-Maddux et al.. 2014). Another study identified elevated
Pb concentrations on rock samples taken near power plant sites, with Pb the most enriched of the 15
elements analyzed by X-ray fluorescence (XRF) spectroscopy (Nowinski et al.. 2012). Pb concentrations
were higher for the skyward facing side of the rock compared with the interior, suggestive of atmospheric
Pb deposition (Nowinski et al.. 2012). In general, Pb concentrations are not correlated with the amount of
electricity generated at an individual plant (Brav et al.. 2017). complicating predictions of Pb emissions
from coal-fired power plants.
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Petroleum-fueled power plants emit -6.4 g Pb/1000 L of fuel oil burned (U.S. EPA. 2006).
Though there are uncertainties surrounding the concentration of Pb in crude oil (U.S. EPA. 2013; Murphy
et al.. 2007). New York City had average Pb concentrations of 3.40 ng/m3 (range 1.22-10.98 ng/m3) in
2009 which land use regression models associated with residual oil burning (Ito et al.. 2016). Used motor
oil, which may be burned in personal space heaters, contains some Pb (Murphy et al.. 2007). Fuel
extraction also contributes to elevated ambient air Pb concentrations. Several studies near the Athabasca
Oil Sands in Alberta, Canada report ambient air Pb concentrations -0.35 ng/m3, though some of this Pb
may come from long-range or regional transport (Granev et al.. 2019; Landis et al.. 2019). and oil fields
of this size are not present in the United States. Biomass fuel consumption has average Pb emissions of 0.
56 mg Pb/kg fuel (U.S. EPA. 2006). Residential wood burning releases airborne Pb at concentrations of
3.3-12.2 mg Pb/kg wood and 2.89-30.3 mg Pb/kg wood for woodstoves and fireplaces, respectively (U.S.
EPA. 2013). Pb-containing particles are ubiquitous in urban areas, indicating widespread emissions from
combustion sources (Murphy et al.. 2007). A mode for Pb urban aerosol was identified at 200 nm, though
Pb was also observed in 50 nm particles, the smallest particle size detected by single particle mass
spectrometry (Murphy et al.. 2007).
1.2.4. Fires
Pb deposited historically in forests is remobilized during wildfires, and Pb in anthropogenic
structures and vehicles can also contribute when wildfires burn infrastructure. Fire emissions account for
18 tons Pb per year according to the 2020 NEI, making wildfires the fourth largest source of Pb emissions
in the United States, behind piston engine aircraft, industrial processes, and fuel combustion, accounting
for -3% of total Pb U.S. Pb emissions (U.S. EPA. 2023a). PM from fires is mostly carbonaceous, but also
contains other elements in low concentrations, including Pb. Preliminary emissions testing results indicate
that more Pb is emitted from smoldering emissions than flaming emissions, and that current emission
factors are substantially lower than previous literature observations, probably because of lower Pb levels
in the environment due to the phase-out of leaded gasoline. Because fires are the largest source of primary
PM in the United States (U.S. EPA. 2021a. 2019). even trace level emissions make fires a potentially
important source of airborne Pb. Limited studies have observed Pb concentrations attributed directly to
smoke from wildfires. Without reliable emissions data before its inclusion in the 2020 NEI, previous
studies focused on air, ash, and soil concentrations to evaluate Pb from fires. This section is therefore
organized as follows: air concentration studies from the previous 2013 Pb ISA, followed by new air
concentration studies, and ending with ash and soil measurement studies. Studies are summarized in
Table 1-2.
The 2013 Pb ISA noted a handful of air concentration studies that have found elevated Pb
concentrations in PMio and PM2.5 during biomass burning episodes (U.S. EPA. 2013). Qureshi et al.
(2006) observed a spike in Pb-PM2 5 at 42 ng/m3 in Queens, NY during a fire event in Quebec. This 24-
hour spike was considerably larger than the 3-month average (July to September) of 5.1 ng/m3. Another
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study quantified the increase of Pb-PMio measured in Finland at 1.7-3.0 times higher during forest fire
emissions from a fire in Russia compared to the reference concentration of 3.5 ng Pb /m3 (Anttila ct al..
2008). Similarly, Golobokova et al. (2020) observed air above Lake Baikal in Siberia before and during
large wildfires. They found levels of Pb were double the base level during the wildfires, with base level
concentrations averaging 0.16 ng/m3 and fire concentrations averaging 0.33 ng/m3. Islev and Taylor
(2020) evaluated trace element and Pb isotope compositions in aerosols from four wildfires near Sydney,
Australia. They found Pb concentrations pre-fire up to -120 ng/m3 and concentrations during and post-
fire up to -210 ng/m3. They attributed 94% of the Pb mass to anthropogenic pollutants, namely historical
Pb from previous emissions. These four studies found elevated Pb measurements, up to 8 times higher,
during days with fire emissions present. The concentrations varied depending on location, with more
isolated locations, such as shipboard on Lake Baikal and rural Finland, measuring lower concentrations
compared with locations with legacy Pb, such as Sydney, Australia and New York City, NY. Despite
these differences, the relative increase is similar across studies.
Also examining Pb attributed to PM2.5, Boaggio et al. (2022) analyzed Pb air concentrations on
smoke-affected days across 13 years. They found Pb to be insignificantly different on smoke days
compared with non-smoke days apart from during wildfires that burned substantial infrastructure. The
median percent change for Pb comparing smoke to non-smoke days was found to be 2-3% lower, but the
maximum was more than 40 times higher at the station that received smoke from the 2018 Camp Fire
which destroyed -18,000 structures. Another study detected Pb-bearing particles in the coarse mode
(PM10-2.5) during the Camp Fire (Sparks and Wagner. 2021). After the burning of the Notre Dame
Cathedral in Paris which contained approximately 460 tons of Pb, an increase in particulate Pb
concentrations from 0.050 to 0.105 (.ig/rn3 was observed 50 km downwind of the fire (van Geen et al..
2020). These results emphasize the importance of accounting for Pb mobilized from burning
infrastructure and vehicles during more destructive wildfires, typically occurring in the wildland-urban
interface.
Pb found from burning of anthropogenic structures was also seen in ash studies. Burton et al.
(2016) looked at ash following a fire in California in 2009 and found Pb was higher in ash samples
collected from burned residences and buildings compared with soil and sediment Pb concentrations.
Similarly, Alexakis (2020) found the median values of Pb in residential ash (78 mg Pb/kg) were 1.5 times
higher than those found in wildland ash (53.5 mg Pb/kg) after a fire in western Attica. The residential ash
sample concentrations of Pb ranged up to 205 mg Pb/kg. Additionally, Campos et al. (2016) studied Pb
levels in burnt soils after a wildfire in Portugal and found unburnt soil concentrations ranging from
approximately 40-50 mg Pb/kg with burnt soil concentrations ranging from approximately 55-150 mg
Pb/kg. These two studies find ash concentrations of Pb similar to those found by Alexakis (2020). who
noted elevated concentrations in residential ash. Soil Pb concentrations ranged from 30-9,000 mg Pb/kg
within 1 km of the Notre Dame Cathedral fire in Paris, whose roof and spire were composed of 460 tons
of Pb (van Geen et al.. 2020). This study also observed elevated concentrations in the direction of
prevailing winds during the fire (van Geen et al.. 2020). The 2013 Pb ISA also cites a study focusing on
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wildfire ash by Odigie and Flegal (2011) that found measurements of Pb in ash following the Jesusita Fire
in 2009 ranging from 4.3 to 51 mg Pb/kg. Another study from the same group measured Pb and other
trace metals remobilized by the Williams Fire in 2012 and found Pb concentration in ash ranging from 7
to 42 mg Pb/kg (Odigie and Flegal. 2014). Both studies concluded the Pb was primarily of anthropogenic
origin remobilized by the fires.
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Table 1-2 Parameters related to fires and associated Pb measurements
discussed herein
Reference
Fire Location
Fire Duration and
Size
Measurement
location
Pb concentration
Qureshi et al. (2006)
Quebec, Canada
July 5-9, 2002
250,000 hectares
Queens, New York
42 ng/m3in PM2.5
Anttila et al. (2008)
St. Petersburg,
Russia
April-August 2006,
unspecified size
Virolahti, Finland
3.5 ng/m3 in PM10
Golobokova et al.
(2020)
Lake Baikal, Russia
July 23-August 1,
2019
Unspecified size
Lake Baikal, Russia
0.33 ng/m3 in
presence of fire,
0.16 ng/m3 without
fire
Islev and Tavlor (2020)
Sydney, Australia
1994, 1997, 2001-
2002, 2004
>1,300,000 hectares
Sydney, Australia
Up to -210 ng/m3
during fire, up to
-120 ng/m3 pre-fire
Boaqaio et al. (2022)
Paradise, California -
Camp Fire
November 8-25,
2018
153,336 acres
Point Reyes,
California
13 ng/m3 in PM2.5
SDarks and Waaner.
(2021)
Paradise, California -
Camp Fire
November 8-25,
2018
153,336 acres
San Francisco,
California
0.26 ng/m3 in PM10
van Geen et al., (2020)
Paris, France - Notre
Dame Cathedral
15 hr on April 15,
2019
6240 m2
50 km downwind in
Paris, France
Up to 0.105 |jg/m3 in
PM10, 30-
9,000 mg/kg in soil
Burton et al. (2016)
Los Angeles,
California - Station
Fire
August 26-October
16, 2009
680 km2
Angeles National
Forest
1.1-16.4 |jg/L in
water
Alexakis (2020)
Attica, Greece -
Attica Wildfire
July 23-26, 2018
1,431 hectares
Attica, Greece
53.5-205 mg/kg in
ash
Campos et al. (2016)
Ermida, Portugal
July 26, 2010
295 hectares
Ermida, Portugal
40-50 mg/kg in
unburned soil, 55-
150 mg/kg in burned
soil
Odiqie and Fleqal
(2011)
Santa Barbara,
California - Jesusita
Fire
May 5-18, 2009
8,733 acres
Santa Barbara,
California
4.3-51 mg/kg in ash
Odiaie and Fleaal
(2011)
Los Angeles,
California - Williams
Fire
September 2-4, 2012
4,192 acres
Los Angeles,
California
7-42 mg/kg in ash
1.2.5. Traffic and Roads
In 2006, the major sources of Pb emissions from on-road mobile sources were fuel combustion
and vehicle wear (U.S. EPA. 2006). After the phase-out of Pb as an anti-knock agent in gasoline for on-
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road automobiles in the 1990s, Pb emissions from tailpipes declined rapidly. As a result, the relative
contribution of non-tailpipe emissions, such as resuspension of Pb in soil and road dust into air as well as
brake, tire, and road wear, has increased. The 2013 Pb ISA found a significant source of Pb in non-
tailpipe emissions from wheel weights. Aucott and Caldarelli (2012) estimated that 13.8 ± 5.0% of the
deposited mass of wheel weights are dispersed each year through abrasion and grinding by traffic (U.S.
EPA. 2013). However, since 2013, wheel weights containing Pb have been banned in many states. In
addition to wheel weights, tire abrasion (mean of two tire samples in Korea =13 mg Pb/kg tire) and brake
wear (30.5 mg Pb/kg brake dust from light duty vehicles in Korea) also contribute to Pb emissions (Jeong
et al.. 2022; U.S. EPA. 2013). When comparing non-exhaust emission sources, asphalt had the highest Pb
concentration (738 mg Pb/kg) followed by road paint (88 mg Pb/kg) (Jeong et al.. 2022). A study that
used material metal concentrations, traffic volume, emissions factors, and sales data to estimate the
quantity of Pb emitted from brake wear and tires in Stockholm, Sweden in 2005 estimated that 24 kg
(0.026 ton) of Pb were emitted from brake wear each year, compared with 2.6 kg (0.0029 ton) of Pb from
tire tread wear; an estimated 549 kg (0.61 ton) was emitted from brake wear in 1998 (U.S. EPA. 2013;
Hiortenkrans et al.. 2007).
Road dust is loose material that can be collected by sweeping and vacuuming the traveled portion
of a road. Also called road sediment or street sediment, it is inclusive of particles associated with non-
tailpipe Pb emissions from traffic (Dietrich et al.. 2022). Road dust also contains PM deposited from other
sources onto or near roads, and is geochemically related to urban soil (Alshettv and Shiva Nagendra.
2022; Dietrich et al.. 2022; Jeong et al.. 2022). Road dust emissions are a function of dust load and
vehicle traffic (frequency of vehicle passing and average weight of vehicles) (Alshettv and Shiva
Nagendra. 2022). with average road sediment concentration of Pb in Busan South Korea in 2014 reported
as 210 mg Pb/kg road dust (Jeong et al.. 2020). Road dust in Philadelphia had mean and median Pb
concentrations of 516 mg Pb/kg and 202 mg Pb/kg, respectively, with higher values reported for
industrial sites and lower for mixed use sites (O'Shea et al.. 2020). while road dust in Toronto had a
median Pb concentration of 63 mg Pb/kg (range 21-220 mg Pb/kg) (Wiseman et al.. 2021). Deocampo et
al. (2012) observed high spatial variability for Pb concentrations in Atlanta road dust, describing a median
road dust Pb concentration of 63 mg Pb/kg in a downtown Atlanta area, but a median of 93 mg Pb/kg and
a maximum concentration of 972 mg Pb/kg in a residential area (also in the urban core). They reported
significant variation on a scale of tens to hundreds of meters. Most road dust particles are large, with sizes
ranging from 10-60 |im (O'Shea et al.. 2021). On average, 13%-18% of road dust analyzed in two areas
in India was <10 |im. 6%-9% was <2.5 |im. and 4%-6% was <1 |im (Alshettv and Shiva Nagendra.
2022). Road dust and soils can serve as both sources and sinks to one another (Dietrich et al.. 2022).
Resuspension of Pb in road dust and soils back to the atmosphere is covered in Section 1.3.4. The
relationship between Pb air concentrations and distance to the road is an emerging area of research. Cahill
et al. (2016) looked at this question for three size fractions of PM in Detroit in 2010. They found that for
coarse PM, Pb concentrations were ~4 ng/m3 at 10 meters from the highway, ~1 ng/m3 at 100 meters
north or south of the highway, and -1.5 ng/m3 300 meters north of the highway. They deduced that the
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increase at 300 meters could be attributed to a heavily trafficked road around 380 meters north of the
highway. For PM2.5 (0.09 to 2.5 |im). they found Pb concentrations were ~4 ng/m3 10 meters north,
~3 ng/m3 100 meters north, and ~2.5 ng/m3 300 meters north of the highway. For very fine PM (0.09 to
0.26 |im). Pb concentrations were -0.75 ng/m3 100 meters south, -0.25 ng/m3 10 meters north,
-0.95 ng/m3 100 meters north, and -0.4 ng/m3 300 meters north. Another study found similar near-road
concentrations for fine Pb with a mean of 5.23 ng/m3 and slightly lower concentrations of coarse Pb with
a mean of 1.14 ng/m3 (Silva et al.. 2021). Contrary to the previous study, they found Pb to be significantly
related to distance to nearest road in coarse concentrations only (Silva et al.. 2021). A third study also did
not find a relationship with distance to road in median PM2.5 water soluble Pb (Oakes et al.. 2016). with a
slight decline with distance for acid soluble PM2.5 Pb and PM10-2.5 Pb. These studies found Pb associated
with PM generally decreases with distance to road. However, the size of this gradient depends on the
particle size distribution of Pb and even with consistent size, there could be subtle differences when
breaking Pb down into water and acid soluble fractions. For a monitoring site in central Los Angeles
located near a major interstate freeway, trends in ambient air Pb concentrations were related to traffic
volume. Pb concentrations decreased slightly from 2005 (median 24-hour sample of .005 (.ig/rn3) to 2013-
2015 but increased from 2015 (median of 0.001 (ig/m3) to 2018 (median of 0.005 (ig/m3). This was
attributed to greater road dust resuspension into air due to increased traffic on the nearby interstate
freeway, since traffic near the site was relatively constant before 2013, but increased considerably from
2013 to 2018 (Farahani et al.. 2021). For context, this increase is substantially less than the overall decline
of airborne Pb concentrations near roads with heavy traffic since leaded gasoline was phased out
(Section 1.4.4). Wiseman et al. (2021) estimated that 90 ± 23 kg of Pb in road dust was resuspended into
air annually in Toronto, Canada, an amount corresponding to 22% of air releases from Toronto industrial
facilities. Limitations identified in this study were identified as uncertainties associated with aging of
street sweeping equipment, street sweeping frequency, and particle size distribution assumptions.
1.2.6. Volcanoes
The 2006 Pb AQCD (U.S. EPA. 2006) included an estimate of 540 to 6000 metric tons per year
for the range of global Pb emissions from volcanoes (Nriagu and Pacvna. 1988). More recently, the
Masaya volcano in Nicaragua was estimated to emit 1 ton of Pb per year (Liotta et al.. 2021). In two
recent studies, Pb concentrations measured at volcanic sources ranged from 0.055 to 0.75 (.ig/rn3 for
samples collected at the main active vent during the 2018 eruption of Kilauea on the island of Hawaii
(Mason et al.. 2021). and 0.14 to 0.27 (.ig/rn3 for samples collected at the rim of a crater of the Masaya
Volcano in April 2000 (Liotta et al.. 2021). Concentrations at the upper ends of these ranges are
comparable to some of the highest currently observed Pb monitoring network concentrations
(Section 1.5). Airborne Pb concentrations associated with the eruption of Kilauea were higher than
concentrations at nearby populated areas (Ilvinskava et al.. 2021). During the week of the 1991 eruption
of Mt. Hudson in southern Chile, observed Pb concentrations more than 2000 miles away on King George
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Island in the Southern Ocean were higher than before or after the eruption (Evangelista et al.. 2022).
Highly elevated Pb concentrations associated with the eruption were also observed in lake sediment
profiles (Evangelista et al.. 2022).
Pb can be emitted in both particulate form (Liotta et al.. 2021). and as a volatile gas at high
temperatures (Edmonds et al.. 2022; Liotta et al.. 2021). and emissions result in the formation of
particulate volcanic plumes downwind of active volcanoes (Edmonds et al.. 2022). Recent research
suggests that emissions of Pb from volcanoes might be underestimated. Ilvinskava et al. (2021) observed
that deposition of Pb and other metals were depleted more rapidly from the volcanic plume of Kilauea
than more widely studied species such as sulfur, and that Pb concentrations in nearby communities during
the 2018 eruption did not change as much as the concentrations of other species. They recognized that
volatile metals like Pb, Cd, and Se were emitted as gases in high temperature volcanic vents and formed
soluble chlorides, sulfates, and sulfides that were rapidly removed by wet deposition in the vicinity of the
source by the rapidly condensing water in the humid environment created by the high abundance of water
vapor emitted from the vent or otherwise present in the humid environment near the source (Ilvinskava et
al.. 2021). They contrasted this with more refractory elements such as magnesium (Mg) and iron (Fe) that
are not emitted as gases and noted that Pb was depleted from the volcanic plume 100 times faster than
these elements (Ilvinskava et al.. 2021). This is consistent with results from Liotta et al. (2021). who
observed enrichment of Pb in rainwater in comparison to volcanic rock at the Masaya Volcano in
Nicaragua. There is some evidence of differences in volatility of Pb emitted from different volcanoes
(Liotta et al.. 2021). Ilvinskava et al. (2021) concluded that emissions of Pb and other volatile metals from
volcanoes, as well as their concentration and deposition in the immediate vicinity of volcanoes might be
underestimated (Ilvinskava et al.. 2021).
1.2.7. Legacy Sources
Contemporary U.S. emissions of airborne Pb described in Sections 1.2.1 through 1.2.6 do not
provide a complete picture of all contributions to ambient air Pb, because Pb emitted from past sources
can become resuspended. Current air emissions are considerably smaller than historical emissions.
Numerous studies of historical records have been reconstructed from sediment and peat cores as well as
long-term soil concentration measurements from many North American locations including Virginia, the
Northeast United States, the St. Lawrence Valley, and northern Alberta (Section 1.3.3.4). Most of these
studies show evidence of decreasing Pb concentrations after the 1970s due to elimination of leaded
gasoline and reductions in industrial emissions (Balascio et al.. 2019; Shotvk et al.. 2016; Sarkar et al..
2015; Richardson et al.. 2014; Pratte et al.. 2013). An exception was a sinkhole near Lake Marion SC,
where sediment Pb concentrations increased continuously during the past 60 years (Edwards et al.. 2016).
This recent research adds to an even larger body of literature summarized in the 2013 Pb ISA and
previous AQCDs that atmospheric Pb concentrations and atmospheric deposition have decreased steadily
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since the 1970s (U.S. EPA. 2013). Pb isotope ratios from some of these studies suggest that historical
sources are an important if not dominant contributor to Pb in North American soil and sediments.
Leaded gasoline has been a major contributor to Pb in the environment, particularly in roadside
and urban soils. An estimated 5.4 million metric tons of Pb additives were used in leaded gasoline in the
United States between 1927 and 1994 (Mielke et al.. 2010). peaking between 1968 and 1972 at more than
200,000 metric tons per year (U.S. EPA. 2013). Pb additive use subsequently declined by 92% from 1970
to 1990 due to health concerns and leaded gasoline was finally banned in the United States in 1996.
Leaded gasoline is a prominent source in sedimentary and other historical records of atmospheric Pb
pollution (U.S. EPA. 2013) and is still relevant near aviation fueling stations. Recent studies investigating
Pb isotope ratios continue to show that leaded gasoline was the principal source of atmospheric Pb (Pratte
et al.. 2013) and the dominant source of Pb in samples of North American sediments (Pratte et al.. 2013)
and forest soils (Richardson et al.. 2014). In the United States, emissions were concentrated in urban
areas, with emissions in 90 urban areas estimated to account for about 30% of total U.S. automotive Pb
emissions in 1982 (Mielke et al.. 2011). In a detailed recent study in a mid-size southern U.S. city, current
roadside soil concentrations decreased since the peak of leaded gasoline usage but remained higher than
geologic background (Wade et al.. 2021).
The United States was a leading producer of Pb in the previous century with major mining and
smelting operations. High Pb concentrations in soils near smelters and other industrial operations have
been observed in numerous studies (U.S. EPA. 2013). The last primary Pb smelter in the United States
closed in Herculaneum, MO in 2013 (Sullivan and Green. 2020). Recent studies also found high Pb
concentrations in soil near closed smelters in Tar Creek, OK; Pueblo, CO; and Eureka, NV (Diawara et
al.. 2018; Chaffee and King. 2014; Zotaetal.. 2011). Pb concentrations in residential soil samples in Tar
Creek, OK ranged from 12 to 2436 mg Pb/kg and averaged 201 mg Pb/kg (Zotaetal.. 2011). and Pb
concentrations in samples from Pueblo, CO ranged from 12 to 10,011 mg Pb/kg and averaged 366 mg
Pb/kg (Diawara et al.. 2018). Pb from closed smelters was also the dominant source of Pb in lake
sediments near Tacoma, WA (Gawel et al.. 2014; Gray et al.. 2013) and attic dust in El Paso, TX (Van
Pelt et al.. 2020). Wang and Kanter (2014) identified 229 former Pb industrial sites in urban areas of the
United States.
Pb-based paint was banned in the United States in 1978. However, 15.3 million homes, 14% of
the homes in the United States, have significantly deteriorated Pb-based paint according to the
Department of Housing and Urban Development's American Healthy Homes Survey. The proportion of
houses with deteriorating Pb-based paint increases with the age of the housing, accounting for 86% of
U.S. houses built before 1940. Regionally, a greater fraction of houses in the Northeast and Midwest
contains Pb-based paint than in the South and West. Housing with Pb-based paint is also unevenly
distributed on a local scale, with a greater fraction of poor and non-white families living in houses with
Pb-based paint (HUD. 2011). A recent county-level geospatial analysis of Pb paint hazard in homes and
childcare facilities found potential Pb hazard hotspots coincided with areas of higher populations of non-
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white children (Baek et al.. 2021). Peeling and deteriorating paint is a source of high Pb concentrations in
yard soil and house dust (HUD. 2011). In streetside, residential, and other soil samples from Durham,
NC, soil Pb concentrations ranged from 6 to 8825 mg Pb/kg, with the highest Pb concentrations observed
within 1 m of pre-1978 residential foundations and both foundation and yard soil Pb concentrations
considerably higher around older houses (Wade et al.. 2021). In urban and industrial areas and near
heavily trafficked roads, historical air emissions together with Pb from deteriorating paint comprise a pool
of legacy Pb in urban soil (Wang et al.. 2022; Obeng-Gvasi et al.. 2021). The potential for the
contribution of legacy Pb to ambient air through suspension and resuspension is discussed in
Section 1.3.4.
1.2.8. Other Sources
Exposure to Pb from community gardens and consumer products is mainly through other media,
but Pb from these sources can briefly become airborne. Pb in food from residential and community
gardens has been the subject of numerous recent studies. Although additional recent research also
indicates that soil Pb can be a concern for urban gardens (Engel-Di Mauro, 2021; Clarke et al., 2015;
Kaiser et al„ 2015; Wiseman et al., 2015), there are ongoing research efforts to improve urban gardening
by reducing Pb contamination in garden produce (Egendorf et al., 2021; Taylor et al., 2021; Gallagher et
al., 2020; Harada et al„ 2019; Fitzstevens et al„ 2017; Brown et al„ 2016; Schwarz et al„ 2016; Spliethoff
et al., 2016; Kaiser et al., 2015).
Pb is also a concern in a variety of consumer products. Batteries were responsible for 92% of Pb
consumed in the United States in 2021 (USGS, 2022). Many of the source oriented Pb monitoring sites in
the national monitoring network for Pb are near secondary smelters for battery recycling (Section 1.4.1).
Recent research has focused on trends for recycling used batteries from the United States and Europe in
other countries. It has been estimated that there are more than 10,000 informal industrial sites for
processing used Pb-acid batteries in low- and middle-income countries, especially in East Asia, South
Asia, and Africa (Ericson et al., 2017). Informal industry is defined as industry characterized by a lack of
adherence to regulation, including zoning and pollution controls (Ericson et al., 2017). In a study of soils
from 15 recycling plants and one battery manufacturing site in 7 countries in Africa, mean soil Pb
concentrations ranged from 480 to 140,000 mg Pb/kg and averaged 23,200 mg Pb/kg inside plant sites,
and ranged up to 48,000 2600 mg Pb/kg and averaged 2600 mg Pb/kg in soil samples from communities
surrounding the plants (Gottesfeld et al., 2018).
By amount of Pb consumed, ammunition ranks second after batteries as an end use for Pb in the
United States (USGS, 2022). The mean estimate of Pb concentrations in soils from shooting ranges in 10
studies was twice as high as Pb concentrations from non-residential Superfund sites (mean values of 3604
ppm Pb and 1868 ppm Pb, respectively) (Frank et al„ 2019). Recent research has advanced our
understanding of the ranges of particle size, solubility, bioaccessibility, and chemical forms of Pb in
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gunshot residue particles from shooting range soils (Schindler et al.. 2021; Sanderson et al.. 2012). There
is a large body of research on the environmental and health consequences of the use of Pb in ammunition
(Arncmo et al.. 2016). Other consumer products that are sources of Pb are contaminated ceramic
cookware, food, toys, cosmetics, antiques, and herbal medicines (Frank et al.. 2019). This summary of Pb
in consumer products is provided for context. Research on health or environmental effects of Pb in
batteries, ammunition, food, and other consumer products is considered beyond the scope of the National
Ambient Air Quality Standards (NAAQS)-related research discussed in this document.
1.3 Fate and Transport of Pb Emitted into Air
Knowledge of Pb transport within and between diverse media, including air, surface water, soil,
and sediment, provides a foundation for understanding the various pathways leading to atmospheric Pb
exposure, as well as the atmospheric contribution to total Pb exposure discussed in Appendix 2
(https://asscssmcnts.cpa.go\ /isa/documcnt/&dcid=359536). Pb emitted into the atmosphere can be
distributed into soil, water, and other media, leading to human and ecosystem contact. Understanding Pb
transport in soil, water, and other media is also necessary for assessing impacts of atmospheric Pb relative
to non-atmospheric sources such as wastewater discharges or mobilization from waste materials.
Sections 1.3.1, 1.3.2, 1.3.3, and 1.3.4 summarize our understanding of fate and transport of Pb in air, soil,
water, and urban media, respectively.
1.3.1. Fate and Transport in Air
Pb is mainly emitted in particulate form, and the fate of airborne Pb is strongly influenced by the
whether it is primarily emitted in the form of ultrafine combustion particles as observed for aviation gas
exhaust, or coarse particles, as observed for resuspended Pb from soil (Section 1.3.4). As described in
Section 1.2.1, Pb is introduced into aviation fuel as the anti-knocking agents tetramethyl and tetraethyl Pb
(Kumar et al.. 2020). During engine operation, the organic functional groups of these compounds are
oxidized and emitted as water and carbon dioxide. A second additive to the fuel mixture, an alkyl bromide
compound, reacts with the Pb present in the combustion mix to form an array of compounds composed of
Pb (II), bromide and chloride ions, molecular ammonia, and other, nonvolatile compounds that form
particles. These particles are either entrained into the engine exhaust or remain in the engine's crankcase
lubricant (NCBI. 2022). Unreacted tetramethyl and tetraethyl Pb have sufficiently high vapor pressures at
ambient and engine operation temperatures to allow for fugitive emissions of these gases (U.S. EPA.
1986). which go on to photolyze in the presence of atmospheric ultraviolet radiation to form Pb
compounds that also contribute to atmospheric PM. These Pb-containing particles are then subject to the
same atmospheric processes that transport and remove other forms of PM. As discussed in
Sections 1.3.1.1 and 1.3.1.2, the transport and deposition of dry particles is defined by size. Depending on
the chemical counter-ion, Pb compounds vary in water solubility, determining the degree to which Pb is
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removed by wet deposition. Figure 1-4 provides a general illustration of the geochemical lifecycle of Pb
derived from fuel additives. Resuspension of soil bound Pb has the potential to contribute to airborne
concentrations near major Pb sources and is considered in Section 1.3.4.
Gas-phase photodegradation
and particle formation
Evaporation
• Bioaccumulation ? j
Runoff
|R3Pb', R;Pb;<|
Dry deposition
Source: Adapted with permission from Encinar and Moldovan (2005).
Figure 1-4 The biogeochemical cycle of tetramethyl/tetraethyl Pb.
1.3.1.1. Atmospheric Transport
The 2013 Pb ISA discussed in detail the available studies concerning the variables governing
long-range and urban-scale transport of particle-bound Pb in the atmosphere, concluding that Pb was
primarily present in submicrometer aerosols, but bimodal size distributions within this size range were
frequently observed (U.S. EPA. 2013). As described in detail in Section 1.2.1, Pb (as PbBr and other
halogenated forms of Pb) is primarily emitted by piston-driven aircraft in the ultrafine particle (UFP) size
range (<100 nm) and larger particles formed from agglomeration of individual particles in the UFP size
range (Turgut et al.. 2020). Particles emitted by aircraft have been observed to be as small as 13 nm in
diameter, when emitted (Griffith. 2020).
Consistent with particles from other sources, according to several studies, particle-bound Pb in
fine PM is transported long distances and found in remote areas. The patterns of dispersion can be
modeled using Gaussian plume models or Lagrangian and Eulerian continental transport models,
indicating that Pb remains a nonvolatile, unreactive particle component. The 2013 Pb ISA also described
studies that indicate that small Pb-containing particles can be scavenged by larger, soil-derived geogenic
particles which can lead to chemical reactions that alter the composition and hygroscopicity of the
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composite particle (U.S. EPA. 2013). This finding was supported by evidence of Pb enrichment of
particles originating from Pb-free sources that were deposited in remote locations.
There has been little recent research on transport of airborne Pb, and recent studies have
continued to focus on transport of Pb associated with particles smaller than 10 (mi. Recent research
supports previous results that Pb derived from high temperature processes such as smelting is largely
emitted in the submicrometer fraction and is capable of being transported over long distances and being
deposited in remote environments (Cullen and McAlister. 2017). On a smaller scale, using a generalized
regression model with 4 km2 sampling grids, Fortuna et al. (2020) demonstrated that Pb content of lichen
samples was significantly spatially associated with dispersion modeling outcomes for Pb, and other
metals primarily associated with PMio emitted from a coal-fired power plant over an area of 176 km2, for
a dispersion model developed for a time frame that corresponds to the average age of biomonitor sample
material. There is also recent research on chemical transformation of Pb. In Beijing, Peng et al. (2020)
reported complete atmospheric transformation of PbO and PbCh from coal combustion to Pb(NC>3)2, from
a process highly dependent on NO2 concentration and relative humidity, and especially efficient at
relative humidity greater than 60% in the presence of sufficient NO2. This observation is important
because insoluble Pb is converted into a more soluble and potentially more bioavailable form (Peng et al..
2020).
Although atmospheric lifetime depends on atmospheric conditions, UFPs quickly grow into fine
particles, or particles smaller than 2.5 |im (PM2.5), by way of gas-to-particle partitioning or coagulation
with other particles before removal. Particles in the size range from 2.5 to 10 (mi (PM10-2.5) and larger are
removed more quickly from the atmosphere than PM2.5 by way of gravitational settling and deposition.
This results in UFP and PM10-2.5 concentrations having substantially greater spatial variability than PM2.5,
with higher atmospheric concentrations typically near their sources and greater spatial variability on
urban and neighborhood scales (U.S. EPA. 2009). More rapid removal processes also result in shorter
atmospheric lifetimes and transport distances for UFP and PM10-2.5 than the PM2.5 size range.
1.3.1.2. Atmospheric Deposition
The 2013 Pb ISA summarized atmospheric deposition patterns for Pb (U.S. EPA. 2013). There
has been a sharply decreasing trend in Pb deposition in the United States and globally since the 1970s,
corresponding to decreasing ambient air concentrations (Section 1.5.1) and decreasing traffic emissions
associated with the removal of Pb from gasoline. In general, fine particulate Pb is mostly soluble and
removed from the atmosphere by wet deposition, and coarse particulate Pb is mostly insoluble and
removed from the atmosphere by dry deposition. Other factors also influence Pb deposition, however.
The pH of precipitation can also play a role because Pb solubility increases with decreasing pH, and
precipitation can also scavenge insoluble particulate Pb as an aqueous suspension. Diurnal variations in
Pb deposition have been observed and attributed to differences in atmospheric structure, specifically
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boundary layer height. Several U.S. studies reported substantially greater deposition rates in areas near
industrial sources than in non-industrial areas, and (U.S. EPA. 2013). As a recent example, regional
differences in Pb deposition patterns have also been documented in a number of studies (U.S. EPA. 2013)
Several recent studies have addressed Pb deposition. Wu et al. (2016) used Pb isotope ratios in
lichens and fungi to show that deposited Pb in bioindicators still reflected historical deposition from
leaded gasoline exhaust. Mazari and Filippelli (2020) focused on urban atmospheric deposition patterns of
Pb and other metals over a wider range of time scales by analyzing soil, bark, and leaves. They oriented
sampling locations along an increasingly urban transect and found the highest Pb levels at the most urban
locations. Previous observations of decreasing Pb deposition with distance from sources also supported
by a recent study showing that previously remediated soil became re-contaminated following aerial
deposition from a Pb smelter, with soil Pb concentrations ranging from 25-100 mg Pb/kg within years of
remediation (Bowers et al.. 2014). Stankwitz et al. (2012) investigated the effect of elevation on Pb
deposition in a forested area of the Northeast United States and found deposition increased with elevation
due to increasing precipitation with elevation. They also found the increase was not linear however,
instead including two abrupt threshold increases associated with the two most common cloud base
altitudes, which in turn corresponded to changes in vegetation. In recent research on cloud processes,
Ebert et al. (2011) observed enrichment of Pb in ice nuclei, with a 25 times higher likelihood of Pb in ice
nuclei than in interstitial aerosols by number in clouds.
Consistent with the 2013 Pb ISA, recent studies continue to show decreasing Pb deposition in
various locations. For example, Perez-Rodriguez et al. (2018) observed a peak in Pb concentrations in
peat samples in southern Greenland that contained Pb transported from North America and Eurasia, and
Wu et al. (2020) used bioindicators to monitor decreasing Pb deposition in Guangzhou after 2000. Other
research continues to evaluate biomonitors for research on atmospheric deposition of Pb and other trace
metals (Kempter et al.. 2017). After deposition, resuspension of Pb in contaminated soil and road dust
into air by traffic, construction, and wind is a potentially important contributor to airborne Pb
(Section 1.3.4).
1.3.2. Fate and Transport in Soil
Knowledge of Pb transport in soil following wet and dry deposition is required to understand risk
and exposure to human and ecological receptors following deposition of atmospheric Pb into soil. The
2013 Pb ISA summarized the long retention time and low mobility of Pb in soil and confirmed the role of
soil as an overall sink for Pb even though atmospheric Pb concentrations peaked several decades ago
(U.S. EPA. 2013). Pb can be deposited onto surface soils in both close proximity to and considerable
distances from point sources. Once deposited in soil, subsequent fate and transport through the soil
column is influenced by several physicochemical factors, including storage in leaf litter, amount, and
decomposition rates of organic matter (OM), composition of organic and inorganic soil constituents,
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mobile colloid abundance and composition, microbial activity, and soil pH. These physicochemical
properties are based on soil forming factors: climate, organisms, parent material, relief, time, and
anthropogenic input. Soils that differ in these factors will subsequently have different physicochemical
properties and different trends in Pb transport. The 2013 Pb ISA summarized studies that describe the role
that each of these physicochemical factors play in Pb fate and transport through soil, and more recent
studies confirm conclusions from the 2013 Pb ISA (U.S. EPA. 2013).
1.3.2.1. Transport into Soil
The 2013 Pb ISA (U.S. EPA. 2013) confirmed findings from the 2006 Pb AQCD (U.S. EPA.
2006) that Pb is deposited from air onto soils in the vicinity of stationary sources and deposition decreases
with increasing distance from the source. As previously discussed, Pb particles of varying size can be
emitted into the atmosphere from several types of stationary sources (Section 1.2) and subsequently
deposited onto soil (Section 1.3.1.2). Pb-derived spatial distribution of contaminants in soils located in the
vicinity of stationary sources, such as non-ferrous smelters, depend on wind direction, size of particles
emitted (smaller particles will travel further than larger particles), and mineralogical and chemical
composition of particles emitted (Ettler. 2016). If soluble forms occur in the dust, greater downward
leaching can occur in the soil profile following deposition (Ettler et al.. 2012). Pb derived from high
temperature processes such as smelting is largely emitted in the submicrometer fraction and is capable of
being transported over long distances and being deposited in remote environments (Section 1.3.1.1). Bing
et al. (2014) demonstrated the long-range transport capability of Pb emissions from industrial sources by
measuring Pb concentrations and isotope ratios in soil profiles from the remote Hailuogou Glacier
foreland in the Eastern Tibetan Plateau. Results revealed Pb enrichment in the O and A horizons relative
to the C horizon, indicting Pb from recent atmospheric deposition rather than parent material. The binary
mixing model using Pb isotope ratios reported that anthropogenic sources contributed to 45.2-61.3 % of
Pb in the O horizon and 8.6-34.8% in the A horizon. Furthermore, the isotopic compositions of Pb and air
mass trajectory models revealed that the major contributions of anthropogenic Pb in surface soils were
from distant sources including Pb ore processing in southwest China and coal combustion in southwest
China and South Asia. The study also discussed potential effects of climate change on soil properties that
would result in Pb release from soils potentially affecting downstream water quality. Binczvcki et al.
(2020) reported similar results, measuring total Pb concentrations and Pb isotope ratios in nine Podzol
profiles located in high elevation remote areas of Kakonosze National Park in Poland. Results revealed
high concentrations of Pb in surface horizons originating from combustion of coal in Poland and the
Czech Republic followed by long-range transport. As described in the 2013 Pb ISA, Pb deposition to soils
has decreased since the phase-out of on-road leaded gasoline (U.S. EPA. 2013). Reduction of Pb surface
soil concentrations since the phase-out are variable, however, particularly in high altitude areas where
there has been little change in O horizon Pb decreases since the phase-out. Kaste et al. (2011) used 21"Pb
measurements to estimate the timescale over which Pb in canopy-derived litter is converted into mobile
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colloid phases that are transported to mineral horizons. Results showed that the Pb is retained in the O
horizon for longer periods of time in areas of higher elevations and latitudes. Similar results have been
reported by Zhou et al. (2019) and Stankwitz et al. (2012). Longer Pb retention times in surface soils at
higher elevations may be due to higher annual precipitation and cloud water depositions as well as slower
OM decomposition due to lower temperatures.
1.3.2.2. Transport within Soil
The 2013 Pb ISA described a variety of complex factors influencing Pb retention and distribution
in soil, including storage in leaf litter, amount and decomposition rates of OM, composition of organic
and inorganic soil constituents, mobile colloid abundance and composition, microbial activity, and pH
(U.S. EPA. 2013). Zhou et al. (2020b) evaluated the Pb adsorption capacity of acidic A and B horizon
mineral soils collected from New York and Vermont. Results revealed that Pb was adsorbed more
strongly in the A horizon than the B horizon soils across all samples indicating the importance of OM in
Pb retention. In addition, soils collected from Vermont were able to selectively adsorb Pb more strongly
than the New York samples. This increase in adsorption was attributed to higher pH, cation exchange
capacity (CEC), manganese (Mn) oxide, non-crystalline Fe oxide, and OM contents in Vermont soils.
The role of leaf litter as both a contributor to Pb in surface soil and as a sink for Pb from soil in
direct contact with leaves was reported in the 2013 Pb ISA (U.S. EPA. 2013). Scheid et al. (2009)
demonstrated that total metals concentrations in leaf litter exposed to manually contaminated soils from
the Swiss Federal Institute for Forest, Snow, and Landscape Research increased over the three-year
duration of the study, suggesting that leaf litter that may come into contact with Pb-contaminated soil
during splashing from rain events can serve as an efficient metal storage pool. Landre et al. (2010)
compared the differences between Pb atmospheric inputs measured in bulk deposition with inputs from
litterfall and throughfall (water depositing onto soil following collection onto leaves) in a remote forested
catchment with limited development in Ontario Canada. Results showed that bulk deposition collectors
may underestimate the amount of Pb reaching the forest floor by about 50%. More recent studies reported
mixed results regarding the role of leaf litter. Luo et al. (2015) measured Pb in soil, litterfall, and plants in
the Gongga Mountain region of Sichuan Province, China and found that both litterfall and atmospheric
deposition were main contributing factors to Pb concentrations in the O horizon. In addition, this study
also revealed a significant correlation between Pb concentrations in fine roots and the A horizon
confirming that fine roots can adsorb and sequester Pb from soil.
After Pb is deposited onto surfaces from litterfall and atmospheric deposition, it is transported
downward as decomposition slowly transforms buried leaf litter into humus. The fate of Pb in litter and
subsequent release to mineral soil horizons occurs over variable timescales that may be strongly
influenced by the rate of organic decomposition. Kaste et al. (2011) used measurements of 21"Pb
throughout soil profiles in coniferous forests in New England and Norway to create a steady-state
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transport model to quantify the fate of metals in leaf litter during OM decomposition over longer time
scales that could be obtained empirically. Results showed the time scale over which canopy-derived litter
was converted into mobile organo-metallic colloids ranged from 60-630 years, varying almost an order of
magnitude, and was slowest in areas where decomposition was slowest. The results of this study also
showed that Pb is retained by the upper litter layer and concentrations increase as litter is buried and
decomposes, resulting in Pb that is enriched in the O horizon. Zhou et al. (2019) reported similar results
in a study that measured Pb concentrations and pools in forest vegetation, litterfall, organic soil, and
mineral soil. In the study, 97.3% of the pools were in litter and organic soil, with Pb concentrations in
organic soil being significantly correlated with total OM in both organic and mineral soil, and
transportation of Pb to mineral soil was dependent on OM decomposition.
Large surface areas with high CEC and negatively charged functional groups make organic and
inorganic soil colloids capable of adsorbing Pb and thus play an important role in Pb transport. Physical
factors influencing colloid mediated transport of heavy metals include flow rate, medium grain size, and
influent concentration (Xie et al.. 2018). Transport of colloids in soil are influenced by flow rate and the
physical and chemical properties of the soil porewater and matrix. Soil porewater with a low ionic
strength and increased colloid and stationary matrix surface charges are associated with colloid
stabilization and maximum Pb-colloid co-migration (Shang and Li. 2011). Conversely, high ionic strength
and lower colloid and surface stationary matrix surface charges are associated with destabilizing colloid
conditions where colloids will tend to coagulate and adsorb onto the stationary matrix (Shang and Li.
2011). When colloids are remobilized from the stationary matrix, Pb that is bound to the colloid
irreversibly is expected to remobilize along with the colloid. However, Pb that is bound via cation
exchange is expected to sorb to the stationary matrix phase following colloid remobilization (Shang and
Li. 2011). Xie et al. (2018) investigated the effects of different colloids on Pb transport under different
physical conditions. Results revealed that compared with deionized (DI) water, montmorillonite, loessial,
and humic acid (HA) colloids all promoted transport of Pb, with HA having the greatest influence on
remobilizing Pb from quartz and sand surfaces with a 33% increase in Pb mobilization through the porous
medium compared with colloid-free DI water. The enhanced mobility was attributed to the large number
of organic functional groups on the surfaces of HA colloids providing large sites for Pb adsorption.
Results also showed that larger matrix grain sizes led to an increase in colloid mobility due to increased
outflow of the colloid in more porous media. Higher flow rate decreased Pb adsorption and colloids on
quartz surfaces, thus increasing mobility of heavy metals and colloids. Shang and Li (2011) studied the
role of rainfall on the migration of Pb-colloid complexes in farmland, floodplain, and Loess platform
soils. Pb migration concentrations with colloids in farmland, Loess platform and floodplain columns were
respectively, 1.12-2.25, 0.91-1.85 and 0.5-2.01 times more than migration concentration with no
colloids. These results confirm the results from Xie et al. (2018). demonstrating that Pb migration is
enhanced by colloid-Pb co-migration.
The 2013 Pb ISA described the effects of microbial activity on Pb sequestration and transport
(U.S. EPA. 2013). Perdrial et al. (2008) observed bacterial Pb sequestration and proposed a mechanism of
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Pb complexation by polyphosphate. They also postulated that bacterial transport of Pb could be important
in subsurface soil environments. Wu et al. (2006) also concluded that Pb adsorption to the bacterial cell
walls may be important with respect to Pb transport in soils. More recent studies suggest that microbial
activity may enhance the release of Pb from both organic and mineral soils. Drozdova et al. (2015)
studied the effects of both live and dead bacteria on the release of trace elements from both organic and
mineral podzols (aqueous solutions in a laboratory system). Results revealed that live bacteria enhanced
the release of Pb to solution, particularly in organic soils, while decreasing the release of potassium (K),
calcium (Ca), strontium, Cu, titanium (Ti), Mn, Zn, and arsenic (As). The authors" noted that decreases in
pH, degradation of dissolved organic carbon (DOC) and metal-organic complexes by microbial activity,
element adsorption at cell surfaces, and biological uptake may occur simultaneously in the soil-bacteria
suspension to both enhance and decrease the release of trace elements from the soil profiles. In the case of
Pb, it is suggested that Pb is released into aqueous solution following bacterial degradation of Pb-organic
complexes.
1.3.2.3. Soil Forming Factors and Land Use
The physicochemical factors influencing Pb retention and distribution throughout the soil column
can vary considerably amongst soils with differences in soil forming factors (i.e., climate, organisms,
parent material, relief, time, and anthropogenic input). The 2013 Pb ISA summarized Pb retention and
distribution through forest soils as strongly influenced by rate of OM decomposition, depth of soil O
horizon, and pH, generally concluding that atmospherically derived Pb will have a longer residence time
in organic surface layers that have lower rates of OM decomposition (U.S. EPA. 2013). Therefore, Pb
will be enriched in the O horizon with increased enrichment occurring in forests where climate, elevation,
and vegetation (i.e., boreal forests versus deciduous forest) favor slower rates of OM decomposition.
Recent literature confirms many of these findings. Richardson et al. (2014) resampled organic and upper
two mineral horizons at 16 sites across deciduous and mixed deciduous/coniferous forests in the northeast
that were previously sampled in 1980, 1990, and 2002. Results revealed that gasoline derived Pb has
leached from the forest floor to mineral soil horizons across the study areas. However, the rate at which
Pb is being transported from the forest floor to mineral soils varied across the 16 sites and was slowest at
sites with frigid soil temperature regimes (STRs) located in the northern portions of the study area. The
decreased Pb response rate and increased retention time in these soils was attributed to slower
decomposition rates in frigid STR and more coniferous vegetation compared with other sites, potentially
decreasing decomposition rates due to higher lignin content. Chrastnv et al. (2012a) compared the
leachability of air pollution control residues in deciduous and coniferous organic soil horizons. Results
revealed higher Pb sorption onto humified OM from coniferous litter compared with deciduous litter. The
increased sorption of Pb in the coniferous organic horizon was attributed to a lower pH and higher portion
of fiilvic acids compared with the deciduous organic horizon, which was a result of differences in
chemical composition and degradability of needles and litter. These results suggest that soil in deciduous
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forest may be more vulnerable to Pb mobilization compared with soils in coniferous forests. Chrastnv et
al. (2012b) compared Pb concentrations and mobility in agricultural and forested soil profiles located at
varying distances from smelting and/or mining release sources. Total Pb concentrations were generally
higher in forested soil profiles compared with agricultural soil profiles. However, Pb in the agricultural
soil profile was found to be more mobile, confirming the important role of forest leaf litter in Pb retention.
Du et al. (2020) investigated the effects of soil freeze thaw cycles (FTCs) on Pb sorption and desorption
behavior in soils vulnerable to alternating periods of freezing and thawing. Results of the study suggested
that FTCs tend to increase Pb immobilization by increasing pH with increasing FTCs, which facilitated
formation of inner and outer sphere complexes. Adsorption capacity was correlated with carbonate and
effects of FTC on Pb adsorption may be more dependent on carbonate and clay content than OM, CEC, or
amorphous Fe.
Burt et al. (2014) investigated and compared the effects of different anthropogenic activities on
trace metal, including Pb, fate and transport. Surface and subsurface soil samples were collected at
locations throughout New York City (NYC) parks (Central Park, Pelam Park, and Van Cortlandt Park)
and from areas in the Bronx Watershed for chemical extraction analysis to investigate and compare trace
element extent, variability, and relationship between soil properties in the two study areas. Central Park
surface samples exhibited higher trace metal concentrations compared with Pelam or Van Cortdlant Park,
which may be related to proximity of Central Park sample sites to public roads and a long history of
intense human activities (shanties, gardening, piggery, and villages) compared with the relatively
undisturbed and mostly wooded Pelam and Van Cortlandt Parks. In the Bronx River Watershed, sum
trace metal concentration was significantly higher in sample locations collected from suburban
Westchester County compared with the more urbanized Bronx. The authors suggested that that the lower
trace element concentrations in the more urbanized area may be attributed to once industrialized land
being recently converted to parkland. Together these results suggest that trace element levels may not
necessarily be dependent on urbanization, current land use, or vegetation, but may be more reflective of
long-term history (type, degree, and age of human disturbances) influencing soil and hydrologic
processes.
Pb concentration trends in NYC parks decreased with depth confirming Pb airborne deposition
from several historical point and non-point sources. Conversely, concentration trends increased with
depth in the Bronx Watershed sample locations. These results were likely a result of soil formation in a
mantle of construction debris covered by anthropogenically transported soil. In addition, the formation of
carbonates from debris materials may have resulted in an increase in pH which increased Pb retention.
The sequential extraction analysis revealed that the predominant forms of Pb were the specifically
sorbed/carbonate-bound (SS/CAR) and the oxide-bound (OX) fractions, indicating that Pb is
predominantly in a form containing the carbonate precipitate, metallic-organic complexes, or metal-
oxides with low bonding forces (i.e., easily mobilized fractions). These results are in good agreement with
a study of NYC garden soils by Cheng et al. (2011) that also suggested anthropogenic Pb was generally in
the highly bioavailable and mobile SS/CAR and OX fractions (i.e., anthropogenic Pb in dust originating
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from urban soils is more toxic and mobile than naturally occurring Pb). The authors found that the
exchangeable and more mobile fractions of Pb were larger in the NYC soil compared with soils found
near a Montana smelter, suggesting that the warmer and humid climate in NYC favored chemical
weathering and trace element mobility. The distribution of Pb in urban soils and the exchange of Pb
between urban soil and other media is further discussed in Section 1.3.4.
1.3.2.4. Summary of Pb Fate and Transport in Soils
In summary, recent literature supports the conclusions from the 2013 Pb ISA (U.S. EPA, 2013)
regarding Pb fate and transport through soils. Studies continue to report higher concentrations of Pb in
soils closer to stationary sources while also demonstrating the potential of Pb being deposited at
considerable distances from sources via long-range transport. Once deposited onto soil, Pb is strongly
retained in organic surface horizons with subsequent Pb retention and distribution in soil strongly
dependent on several physicochemical properties, including storage in leaf litter, amount and
decomposition rates of OM, composition of organic and inorganic soil constituents, mobile colloid
abundance and composition, microbial activity, and pH. In general, leaf litter, low rates of OM
decomposition, neutral pH, and soil constituents rich in charged surfaces such as OM, Fe and Mn oxides,
and clay minerals will lead to increased Pb retention and sorption. Conversely, thin organic layers,
increased OM decomposition, acidic pH, increases in anthropogenic Pb, and less reactive soil constituents
such as quartz will tend to increase Pb leaching from soils.
1.3.3. Fate and Transport in Water and Sediments
As discussed in the 2006 Pb AQCD and the 2013 Pb ISA (U.S. EPA. 2013. 2006) atmospheric
deposition, urban runoff, and industrial discharge are large contributors of Pb to surface waters with
greater runoff being linked to larger storm events following a dry period. Evidence from the 2013 Pb ISA
also found some evidence of a seasonal effect on runoff with greater runoff following snowmelt or rain-
on-snow events. Pb transport and sedimentation within aquatic systems depends upon chemical properties
including water pH and salinity, presence of OM and iron (Fe) and Mn oxides, total suspended solids
(TSS), as well as mechanical processes including turbidity and flow. Previous research has also shown Pb
is relatively stable in lacustrine and riverine sediments but resuspension of sediment into water or
dissolution from sediment often occurs during and following storm events and can be a larger source of
Pb to the water column and downstream reaches than concurrent atmospheric deposition. New research
primarily provides additional support for the 2006 Pb AQCD and 2013 Pb ISA (U.S. EPA. 2013. 2006)
conclusions with additional information on runoff following fire events and seasonality influence on
transport and sedimentation. Furthermore, new literature provided information on temporal trends of Pb
concentrations in sediments and several studies are summarized in Section 1.3.3.4 to highlight the
importance of legacy Pb pools as potential "new" sources of Pb to waterways.
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1.3.3.1. Biogeochemistry
1.3.3.1.1. Freshwater Biogeochemical Influences
The transport of Pb through freshwater systems is influenced by a variety of biogeochemical
factors such as OM content, redox, alkalinity, and seasonality. Since the 2013 Pb ISA (U.S. EPA. 2013).
new information was found for how Pb transport and availability is increased in the presence of higher
nutrient levels and under anoxic conditions, while photolysis of OM reduces Pb concentration because it
can be bound more to organic molecules. There is also an improved understanding of the mechanisms for
how different types of OM (e.g., HAs, or amount of aromaticity) interact with Pb and how dissolved OM
and PM can increase the mobility and solubility of metals in aquatic systems. An increase in DOC leads
to a decrease in the amount of Pb bound to PM because Pb instead binds more to dissolved organic matter
(DOM). Thus, activities that increase DOM (like surface mining or heavy rain events) can increase the
mobility and solubility of metals in aquatic systems (Gueguen et al.. 2011). Similarly, Chen et al. (2019)
found that the solubility and mobilization of Pb increases through the formation of Pb-DOM complexes.
As the DOMs become more allochthonous, more humic-like, more aromatic, and optically darker, the
active Pb-binding fraction increases (Chen et al.. 2018). Coordination chemistry has shown that Pb
predominantly binds to the phenolic and carboxylic group on a salicylic-type structure or two adjacent
carboxylic groups on catechol-type structures. Cabaniss (2011) found Pb(II) preferentially binds to strong
amine-containing sites which are often located on small molecular weight (MW < 1000), and lower
aromaticity molecules. There are no highly aromatic molecules acting as strong ligands for Pb(II). Pb(II)
binds to phenolic groups but is more strongly bound by amine groups. At low metal loading, Pb(II) is
selectively bound to very high molecular weight compounds (>2000 amu). And Pb(II) are bound by
molecules with less negative overall charge (average charge of occupied ligand molecule Zbnd > —1.6) at
relatively low metal loading. At pH 7.0, Pb(II) binds in the order amines > phenols > carboxylates.
Jeremiason et al. (2018) found that DOM mobilizes historical deposits of Pb from bog peatlands and Pb
and DOC concentrations were correlated in the bog. The key factor is DOM leaching or production
leading to Pb redistribution among binding sites in solid peat and the dissolved phase. The amount of Pb
mobilized per unit DOC (|ig/mg DOC), was greater in bogwater (0.047; range 0.037-0.067 at individual
sites) than lagg water (0.033; range 0.007-0.050). Interestingly, Pb was found to preferentially adsorb
onto bacterial cells (organic material) than on clay minerals (Du et al.. 2016).
Suspended particulate matter (SPM) and sorption material also influence Pb transport and
availability. SPM significantly influences Pb adsorption. Total Pb concentrations in water were higher
when the content of the PM in the river water was high (Milacic et al.. 2017). Further, the highest
partitioning coefficients observed for Pb were a consequence of its high affinity to SPM and low Pb
solubility in water. Pb binding to SPM in the lower Waikato River in New Zealand is predominantly via
Fe-oxide surfaces and can be reliably predicted using surface complexation adsorption modeling
(Webster-Brown et al.. 2012).
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Other metals, nutrients, and inorganic compounds in the sediment and open waters can affect Pb
mobilization. High nutrient levels can increase the potentially mobile fractions of Pb (Kang et al.. 2019).
A significant negative relationship exists between total phosphorus (TP) and Pb concentrations per unit
mass of phytoplankton in lakes (Gormlcv-Gallaghcr et al.. 2016). Sulfide also showed a negative
relationship with Pb, likely reflecting precipitation of Pb-sulfide complexes in sulfide-rich porewater
(Carling et al.. 2013). Lombardi et al. (2014) found that the percent labile Pb (86 %) compared with
percent dissolved Pb suggests that most of the Pb was complexed with inorganic compounds. Pb was
complexed preferentially with COr (25 %), NO;, (22 %), and OH (19 %). Chlorophyll a and TSS were
also correlated with most Pb fractions. Groundwater may be contaminated with Pb, and this may be due
to strong correlations between Pb and Fe or Mn oxides and with total dissolved solids (Wang et al.. 2016).
and this study suggests that groundwater contaminant of Pb is due to natural processes and not from
surface water contamination.
Pb transport and availability are influenced by redox conditions. Pb is typically released from
sediments in anoxic environments and adsorbed from the overlying water in an aerobic environment
(Kang et al.. 2019). For the exchangeable fraction (characterized by soluble species, species with cation
exchanges sites, and carbonate-bound species), Pb increased in aerobic conditions and for both high and
low nutrient levels but decreased under anoxic conditions. Sediment absorbs more Pb2+ under aerobic
conditions. Another study found that Pb can precipitate under either oxic or anoxic conditions, but due to
different mechanisms. In a eutrophic lake, Chen et al. (2019) observed that algae degradation may
decrease redox state in sediments and sulfide may be released from the degraded algae, which may
promote the formation of Pb-sulfide precipitations under anoxic conditions. Whereas in oxic conditions
(high redox state), Fe2+ and Mn2+ oxidation may occur in sediments and result in the adsorption or
coprecipitation of Pb. Also, Pb can be confined to an immobile form (organic sulfides) at higher alkalinity
in stream sediments, and gradually be released due to chemical weathering (Pearson et al.. 2019). Li et al.
(2016) found that in acid mine drainage, Pb was found to be associated with the carbonate fraction and
under waterlogged conditions, dissolved Pb increased when Fe increased in concentration. In waters with
acid mine drainage, Pb was predominantly present in the residual (77.7%-85%), followed by oxidizable
(9.4%-12%) and reducible (5%—10%) fractions. Also, the decomposition of OM like cyanobacteria can
cause a reduction in the oxidation reduction potential (ORP), which can result in an increase in Pb bound
to sulfate ions (Ni et al.. 2019). Chen et al. (2019) also found that Pb in both eutrophic and non-eutrophic
lakes was commonly complexed as Pb(HS)2, PbCO, and to a lesser extent: Pb2+, PbOH+, Pb(OH)2,
PbS04, Pb(C02)2, and PbHC03+.
Temperature and seasonality also influence Pb adsorption and transport in freshwater systems. Pb
showed higher concentrations during the spring than summer in river samples (Zhang et al.. 2016a).
Zhang et al. (2016a) also found that in the spring, the majority of Pb is found in the inert form (not
reactive with NH4+ or OH" ion exchange resins) and only ~1 to 5 % of Pb is found in the organic or labile
forms. But in the summer, there were higher percentages of labile Pb ranging from ~10 to 60% in the
rivers. The organic fraction was the same in both seasons, while the labile fraction increased (on average)
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from 6.75 to 19.95 % between spring and summer. On average, the labile Pb fraction increased in all the
rivers during summer. The increase in labile concentrations might be attributed to human activities,
leading to increased potential toxicity in these rivers. In winter months, Sun et al. (2018) found that Pb2+
has a low binding energy in ice compared with other cations (Fe > Cu > Mn > Zn > Cd > Hg > Pb).
Lombardi et al. (2014) found that total, dissolved, complexed, and labile Pb species were all higher in the
winter, while Pb was present more in the particulate form in the summer. Chen et al. (2019) observed the
highest dissolved Pb concentrations in July for a eutrophic lake, while the highest dissolved Pb
concentrations were in January for a lake covered by floating and submersed macrophytes. The greatest
increase in Pb complexation with DOM occurred in the eutrophic lake in July, while it occurred in the
non-eutrophic lake in April. However, the degree of Pb complexation to DOM was significantly larger in
the non-eutrophic lake in all seasons. The mobility of Pb in sediments showed significant seasonal
variations, reflected by a high release of Pb during the spring and summer in the algae-dominated region
and during the autumn and winter in the macrophyte-dominated region. A possible mechanism is that in
the algae-dominated regions of the lakes, increased bacterial abundance in the sediments during the spring
promoted microbial reduction of Fe/Mn oxides, which likely released Pb from sediments.
Pb was found to be higher during periods of extreme flooding (Milacic et al.. 2017). where Pb
inputs are primarily derived from heavy industrial activities or mining and metallurgy activities
(depending on the site). Valencia-Avellan et al. (2017) also found that Pb concentrations increased with
peak flow in an ephemeral tributary. This is likely because Pb is strongly associated with both particulate
and colloidal Fe and Al oxides, and also cerussite (PbCOs), SO4, and DOC, which can all increase during
high flood periods which are associated with the resuspension of sediments into water.
Light also has an influence on the dynamics of Pb in freshwater systems. Within an acid mine
drainage impacted wetland, it was found that increased light levels caused a reduction in ferrous iron and
this was associated with an increase in Pb concentration (Duren and McKnight. 2013). The mechanism
for this process is the formation of superoxide radicals (O2) and H2O2 from the photoreduction of DOM
in the wetlands, where H2O2 reacts with Fe2+ and converts it to Fe3+ (reducing the amount of Fe2+ during
the day). During photolysis, Drozdova et al. (2020) observe two simultaneously occurring processes: (1)
the degradation of high molecular weight organo-mineral colloids and the formation of low molecular
weight organic molecules and Pb complexes, and (2) the formation of the >0.22 mm particulate
aggregates of Pb and OM. The DOM degradation produces both CO2 and HCO3" whereas Pb which is
initially associated with organo-ferric colloids are subjected to coprecipitation with newly formed Fe(III)
oxy(hydr)oxides. Also, photolysis caused a decrease in Pb by 48% in solution and this may be because Pb
is correlated to changes in concentration of Fe, DOC, and humic substances. For instance, Fe-OH and
organic ligands can form ternary surface complexes with Pb. An alternative mechanism of metals removal
could be their precipitation in the form of individual metal hydroxides that occurs after photo-degradation
of metal-ligand (Me-L) complexes.
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1.3.3.1.2. Saltwater Biogeochemical Influences
The transport of Pb through saltwater systems is influenced by a variety of biogeochemical
factors such as salinity, organic matter content, redox, alkalinity, and seasonality. Since the 2013 Pb ISA
(U.S. EPA, 2013), new information was found on how Pb concentrations in solution increases with
increasing salinity and temperature, but decreases in the presence of DOC, Fe(III) and Mn(IV/III)
(hydr)oxides, which provide important binding sites for heavy metals under high dissolved oxygen (DO;
oxic) conditions. There is also more information on the role of sulfide in estuarine systems and on
whether Pb comes from anthropogenic or natural sources.
Salinity of estuarine and coastal waters can have a strong influence on Pb fate and transport. (Yao
et al„ 2016) found that the concentration of Pb adsorbed to PM decreases with increasing salinity in the
medium-low salinity of the estuary near the river mouth, indicating the release of Pb during early mixing
stages in the estuary. The metal release resulted from a balance between two opposite processes: (1) metal
mobilization due to ionic exchange or degradation of organic complexes and (2) metal re-adsorption onto
an existing or newly formed solid phase. Basically, with increasing salinity, cations such as Na+, Ca2+,
and Mg2+ compete for the adsorption sites on particle surfaces, thereby decreasing adsorption and
enhancing the release of sorbed Pb from the particle surfaces. Similarly, Zhao et al. (2013) observed that
as salinity in the Yangtze Estuary increased, Pb was released from the sediments, but it was minimal
(0.004-0.017%), due to preferential retention in Fe-Mn oxides and organic content. Pb cations seemed to
be sorbed more specifically to sites with high dissociation constants (and high sorption energies), making
them less vulnerable to leaching. (Karbassi et al., 2014) also observed greater Pb flocculation at lower
salinities (0.5%) and constant pH of 8. However, another study found the opposite pattern; while the
partition coefficient Kd (L/kg), which is the relationship between the sorbed state to the dissolved state of
a metal, generally decreases with salinity due to higher ionic strength and competition for sorption sites,
Alkhatib et al. (2015) found the Kd for Pb increased with salinity (with Kd values at 234 L/kg in
freshwater, 575 L/kg in brackish water, and 1341 L/kg in seawater) and this is mainly attributed to
formation of insoluble metal species, like PbS04, which led to higher Kd values with the increase of
salinity. Another study with this pattern found that for PbS, freshwater exhibited the highest Pb release
followed by seawater and estuary water (Chou et al., 2018). Clearly, other factors like the types of
inorganic species and metals and other conditions must influence the impact that salinity has on the
transport of Pb in saltwaters.
The presence of different minerals in estuarine and seawaters can influence the transport of Pb
between sediments, saltwater, and the atmosphere. For example, Shelley et al. (2018) observed that Pb
and Al was significantly correlated (r2 = 0.478) in saltwater and that Pb solubility was greater in saltwater
than in ultra-high purity water used as a control, though Pb solubilities decrease as aerosol loading
increased. Fe(III) and Mn(IV/III) (hydr)oxides provide important binding sites for heavy metals under
oxic conditions, and sulfide provides important binding sites for Pb under anoxic conditions (Wang et al.,
2013). Consequently, the reductive dissolution of Fe(III) and Mn (IV/III) (hydr)oxides could encourage
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the release of Pb into solution. But dissolved levels of Pb became undetectable within 10 days suggesting
that it can be almost completely sequestered in the metal sulfide phases under sulfate-reducing conditions
(during bacterial sulfate reduction activity). Morgan et al. (2012) found acid volatile sulfide to have a
strong relationship with reactive Pb in estuaries and a strong relationship between FeS and Pb in
sediments, where Pb sulfates are more likely to precipitate than FeS due to lower solubility. Thus, FeS is
likely to retain Pb in estuarine sediments. Tovar-Sanchez et al. (2019) observed that both Pb and Fe were
abundant in the sea surface microlayer. And Keene et al. (2014) saw a strong correlation between total Pb
and reactive Fe in interfacial sediments of an estuary, with 50% of Pb being associated with reactive or
"acid extractable" phases in the sediment. Ebling and Landing (2015) studied the Pb levels in the sea
surface of the open ocean ("microlayer" - the thin layer at the boundary between the ocean and the
atmosphere) and measured dissolved, labile particulate, and refractory particulate trace element
concentrations of the sea surface microlayer. They found dissolved Pb to increase in the microlayer by a
factor of 2-3 over time, coinciding with an increase in Fe, which may have come from precipitation. At
the same time, the refractory particulate Pb increased by a factor of 23 in the microlayer. Pb in the
microlayer had retention time of about 1-2 days. The enrichment factor (EF) for Pb was >1 demonstrating
enrichment in the microlayer. Canovas et al. (2020) observed Pb was abundant in PM (37%-59% in
dissolved fraction) and that 66% of Pb was found forming CI" complexes, -20% as CO3" complex, 5% as
Free Pb, and 5% as Pb hydroxide. Pb showed a balanced speciation between the uncharged and positively
charged species.
PM suspended in water has been found to strongly influence Pb transport and fractionation in
seawater. Angel et al. (2016) reported that the amount of dissolved Pb concentrations in seawater was
dependent on the concentration of precipitate present, decreasing as the precipitate concentration
increased. The composition of the precipitate formed is likely to be a metastable Pb chlorocarbonate.
Feng et al. (2017) observed how the partitioning coefficient (Kp), for the amount of Pb sorbed to SPM
was highest for Pb compared with other trace metals (Ni, Cr, Cu, Hg, Zn, Cd, and As). And the Kp for Pb
is higher in the SPM than for the sediment-water interface. This is because SPM has a smaller particle
size, and higher specific surface area and OM content, and thus can adsorb more heavy metals. The
exchangeable and carbonate fractions of Pb also had significant positive correlations with the Kp for Pb in
SPM and the exchangeable, carbonate, and residual fractions of Pb had significant positive correlations
with Kp for Pb in sediments. Thus, adsorption is likely to be the dominant partition process of Pb. Burton
et al. (2019) found through a review study that the removal efficiencies (of the metal from the water
column to SPM) for 7 of the 12 estuaries were at or greater than approximately 75% for Pb. And metal
removal efficiency was greater for Pb than Cd and Zn, consistent with the metal's partition coefficient. Pb
accumulates more in the finer fractions of clay (<8 |im) and fine silt (8-16 |im) (Yao et al.. 2016). Pb
concentrations in the bulk SPM varied from 25 to 38 mg Pb/kg, with an average of 32 mg Pb/kg. Pb had
an average EF value of 0.81 and were all <1.5. This indicates that the Pb concentrations originated from
natural weathering processes. The EF of an element is defined as the ratio of that element to a
conservative element in a sample divided by the ratio of that element to the same conservative element in
a background reference sample. An EF value between 0.5 and 1.5 suggests that the trace metals may be
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entirely from crustal materials or from natural weathering processes, while an EF value >1.5 suggests that
a significant portion of trace metals is delivered from non-crustal materials, or nonnatural weathering
processes, like anthropogenic activities (EF = [Pb]/[Fe]sample / [Pb]/[Fe] (Yao et al.. 2016). In very
polluted environments, extreme EF values for Pb can be observed. For example, an EF of approximately
600 was recently observed in a highly industrialized and urbanized area of China, indicating Pb that was
dominated by anthropogenic contributions (Xing et al.. 2017). One study by Holmes et al. (2014)
examined the role of estuaries in modifying the adsorptive properties of new and aged plastics towards
trace metals and found the absorption capacity of Pb on plastic surfaces to decrease from river water to
seawater and with decreasing pH due greater competition with other cations.
Temperature, oxic conditions, and organic content influence Pb transport in saltwaters. Within
estuarine waters, temperature was positively associated with free Pb concentration, with a 1°C increase
corresponding to approximately a 7% increase in free Pb concentration (Dong et al.. 2016). DO was also
found to be dominant factor that controlled the release of Pb from coastal sediments, with increased
hypoxia causing increased Pb in overlying waters compared with sediments (Liu et al.. 2019). Similarly,
Banks et al. (2012) observed greater dissolved Pb concentration in porewater estuarine sediments at lower
DO levels, where the ratio of dissolved Pb concentration to metal concentration was 1.2 for 5% DO, 1.1
for 20% DO, and 0.9 for 75% DO. In anoxic conditions the presence of wetland plants, like S.
alterniflora, could lead to higher concentrations of Pb in the sediments, via pumping oxygen into the
rhizosphere, which can cause the release of Pb to sulfates (Wang et al.. 2013). Similar to freshwater
systems, within estuarine waters, DOC and free Pb concentration had a negative relationship, indicating
organic ligands in the water column were more important binding agents for free Pb ions relative to
particulate organic ligands (Dong et al.. 2016). Also, the presence of HA showed inhibition effects on its
metal release of Pb from PbS (Chou et al.. 2018). Carbon dioxide (CO2) also influences Pb; a model
predicted that under scenarios of increasing CO2, free Pb could increase from 9%-97% and organically
bound could increase by 5%-43% (Stockdale et al.. 2016).
Several studies also found that within saltwater systems, pH has a negative impact on free Pb
concentrations. Vasvukova et al. (2012) also found that % dissolved Pb decreases as pH increases, and Pb
is an element strongly associated with colloids and exhibits significant increases of relative proportion of
colloidal forms with pH increase. Another study found that the partition coefficient Kd (L/kg), which is
the relationship between the sorbed state to the dissolved state of a metal, was greatest for Pb at pH 7, at
16434 L/kg, indicating that more sorbed Pb was present at neutral pH (Alkhatib et al.. 2015). When
assessing the precipitation of Pb from PbS, there was minimal release of Pb from PbS at pH 8,
intermediate Pb released at pH 7 (213 mg/L/m2) and even more Pb released at pH 5 (386 mg/L/m2), which
suggests that H+plays a role in the oxidative dissolution of Pb sulfides (Chou et al.. 2018).
Seasonality, rainfall, and tidal flows can influence Pb dynamics in estuaries and coastal waters. In
the study by Hierro et al. (2014). Pb was found primarily in PM (average of 825 (.ig/L). Pb-PM
concentrations (per volume) were 2-3 times higher in PM carried by ebbing tide compared with the rising
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tide, due to increased PM when there is a fall in sea level. Often, Pb sorbed/coprecipitated with Fe
hydroxides, and highest particulate concentrations coincided with the estuarine maximum turbidity zone.
In terms of rainfall, one study found that increased rainfall resulted in lower free Pb concentrations, likely
due to dilution (Dong et al.. 2016). while another study found higher levels of Pb in the spring when the
rainfall amounts where larger compared with summer months (Xing et al.. 2017). Additionally, this study
found strong correlations between crustal-derived elements (Al, Fe) and anthropogenic elements (Pb, Cd,
Zn) likely due to both being influenced by air deposition rainwater runoff (Xing et al.. 2017). Also, during
seawater-freshwater interaction from seawater intrusion to an aquifer, it was observed that Pb exhibited
significant correlation with colloids and was thus sensitive to the flow of the colloidal fraction where
seawater and freshwater are interacting (Tan et al.. 2017).
1.3.3.2. Transport into Water (including Runoff)
The 2006 Pb AQCD concluded Pb in runoff was mostly in the particulate fraction and identified
runoff as being dependent on storm intensity and time between rain events (U.S. EPA. 2006). The 2013
Pb ISA provided information on Pb runoff from roadways, urban areas, and snow melt into watersheds
(U.S. EPA. 2013). New research provides additional support for both the 2006 Pb AQCD and 2013 Pb
ISA conclusions with additional information on runoff following fire events and urban sources of Pb
unique to city history and planning.
1.3.3.2.1. Urban
Pb in runoff in urban areas is correlated with surrounding land use characteristics such as
impervious surface area and road density. In a study that examined Pb concentration and distribution in
bed sediments of the Palolo drainage basin in Hawaii, Hotton and Sutherland (2016) found of the three
streams that comprise the basin, Palolo had Pb concentrations of 134 mg Pb/kg compared with Pukele
(24 mg Pb/kg) and Waiomao (7 mg Pb/kg). Furthermore, Palolo had high concentrations along its entire
length whereas Pukele and Waiomao showed highest concentrations downstream. This high Pb
concentration along Palolo was correlated with urban land-use characteristics including street length,
storm drain length, the number of storm drain inlets and outlets, as well as vehicle counts and overall
population. Urban development around Palolo was higher than around the other two streams in the
drainage basin. In a similar study, sediment Pb concentration was measured in highly urbanized
watersheds across several U.S. cities (Nowell et al.. 2013). Pb was positively correlated with local urban
factors and study area variables including population density, urban land cover, road density, and amount
of impervious surface area as well as total organic carbon. Boston had higher Pb concentrations than other
cities with comparable urbanization and sediment total organic carbon and the authors highlight this
higher Pb concentration in Boston likely reflects the city's long history of industrial activity and high-
density development. McKenzie and Young (2013) examined Pb water column fractions following storm
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events in creeks draining different surrounding land-use areas. Highways and urban areas had higher
runoff loads compared with agricultural and natural sites. Agricultural storm loadings were similar to
those in natural systems and irrigation loadings were less than storm loadings. Pb was primarily
associated with suspended sediments so would have low mobility and bioavailability. Highway runoff, on
the other hand, had high levels of dissolved Pb.
Several recent studies highlight urban-specific sources of Pb to city waterways due to city design
and historical pollution. For example, a study by Coxon et al. (2016) examined and mapped the
contamination histories of the rivers that drain the lower western Chesapeake Bay basin. Sources of
contamination have changed over Virginia's history and reflect the development of the area. Western
mountain reaches have elevated Pb levels due to lithology and historical mining while agriculture and
urbanization contribute to Pb enrichment across the drainage basin. Norfolk naval base and shipyard is a
current and significant source of metal enrichment, as are incinerators, older office buildings with Pb
paint, and ordnance storage. Furthermore, changes in urban land-use management have led to legacy Pb
pools becoming a new source and Pb enrichment downstream of urban areas is high—with sediments
downstream of Richmond showing particularly high Pb enrichment levels. Overall, fuel combustion,
street dust, and highly contaminated urban soils are the contemporary suppliers of Pb to the Virginia
Chesapeake waterways. In the Gwynns Falls watershed area in Baltimore, Pb concentration in riparian
sediments decreased with increasing distance from the city center (when normalized for sediment surface
area). Also of note, three hotspots of contamination in the urban system occurred adjacent to areas that
had been identified in 1979 as artificially filled (Bain et al.. 2012). A non-U.S. study found that despite
restoration efforts enacted 30 years prior, Pb sediment contamination in an urban lagoon in Sardinia
exceeded 100 mg Pb/kg at 22 of 34 monitoring locations (Atzori et al.. 2018). High Pb concentrations
were not correlated with OM content and showed similar contamination patterns with mercury (Hg).
Instead, Pb (and Hg) peaks were in sites with proximity to a chlor-alkali plant and an airport. In a study
that examined trace metal export in relation to soil concentrations, soil and water properties, and
watershed land use across several New England watersheds, Pb export rates varied from 0.03 to 0.37 kg
Pb/year/km2 (Richardson. 2021). Dissolved Pb concentration was not correlated with soil Pb
concentration but was positively correlated with aquatic Zn and DOC concentrations. Furthermore,
dissolved Pb export was positively correlated with watershed cover of wetlands and negatively correlated
with the percentage of forest cover. The author suggests the positive correlation with wetland cover may
be due to wetlands serving as a reservoir of historic pollution and the negative correlation with forests is
due to there being less development and therefore less pollution, as well as potentially higher retention of
trace metals (Cu, Pb, and Zn) by soil iron oxides. Another recent study sought to assess the trace metal
loading rates in the Great Lakes basin and estimated results found Pb inputs to Lakes Superior and
Michigan were primarily atmospheric while Lakes Erie and Ontario received proportionally more Pb
from tributary inputs and to a less degree, from connecting lake channels. Lake Huron, being in the
middle, unsurprisingly receives Pb from atmospheric deposition and tributaries in relatively more
equivalent contributions (Bentlev et al.. 2022). Lastly, another study examined the effects of road salts
and deicers on metal mobilization within the soil profile (Schuler and Relvea. 2018). Sodium (Na) can
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displace Ca and Mg in the soil which can increase porosity of the soil structure leading to mobilization of
metals. Salts can also mobilize Pb by breaking down the organic-rich colloidal structures that often bind
Pb within the soil matrix. Salts can also displace metals, including Pb, from binding with organic
compounds because the binding affinity is higher for Na, Ca, and Mg than it is for heavy metals.
Although this study did not measure Pb concentration in water or stream sediments, it highlights how
salts and road deicers can increase Pb mobility into water systems through the physical and chemical
changes to the soil matrix—an interaction effect unique yet widespread in urban systems.
1.3.3.2.2. Non-Urban
As in urban environments, Pb in runoff in non-urban areas is primarily within sediments and
runoff into waterways is driven by storm events and overall precipitation patterns. A European study
examined trace metal budgets across 14 forested catchments over the period of 1997-2011 (Bringmark et
al.. 2013). Due to high anthropogenic deposition for decades, Pb accumulation in catchment soils was
high. Pb is bound to soil OM in these soils leading to high retention of Pb in the system. At higher altitude
sites which experience greater precipitation, retention is lower due to greater runoff and transport of PM
out of the system. In a similar study in England, while Pb deposition has decreased in recent decades,
legacy Pb in peat catchments is a continuing source of Pb to waterways (Rothwell et al.. 2011).
Atmospheric deposition was measured at 34 gPb/ha/year while fluvial outputs were 316 gPb/ha/year.
Following storm events, Pb runoff into waters occurs primarily as suspended particles (261 gPb/ha/year)
with a smaller aqueous portion (55 gPb/ha/year). In the Snowy Mountains of Southeast Australia, down-
catchment reservoirs with large catchment size showed comparable sediment metal enrichment values to
soil enrichment values indicating the soil surface which contains metal values reflective of anthropogenic
deposition is the source of sediment to reservoirs, not eroded subsoils (which is less contaminated)
(Stromsoe et al.. 2015). However, Pb (and chromium) are depleted in reservoir sediments compared with
soils, indicating these more particle-reactive materials are bound up within the soil matrix and are not
being washed into down-catchment reservoirs.
In waterways draining the Alberta badlands, total Pb concentration frequently exceeded Alberta
guidelines for freshwater biota (Kerr and Cooke. 2017). However, Pb concentration was positively
correlated with TSS due to an increase in sediment mass, not due to increased sediment Pb concentration,
highlighting the importance of erosion and precipitation interactions in arid systems. In contrast, a study
in Brazil found metal fluxes, including Pb, were highest during the dry season compared with the wet
season and were due to suspended sediments (Bezerra da Silva et al.. 2015). Unlike the studies measuring
flux in the badlands of Alberta, this increase was not due to bedload but due to a lack of dilution. The
watershed included agricultural and industrial sources as inputs and Pb was likely from coal combustion,
solid waste incineration, and legacy Pb from petroleum but sources could also include pesticide use,
sewage sludge runoff and agricultural wastewater.
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Pb runoff and accumulation from soils into lakes is also influenced by snow and ice melt. Overall,
three catchments in the Pyrenees have greater Pb concentration in lake sediments than surrounding soil
and bedrock (Bacardit et al.. 2012). Lakes within the two watersheds above 2000 meters above sea level
had lower Pb concentrations than lakes within the lowest watershed (1655 meters). The lower Pb
accumulation in these high elevation lakes may be due to snowmelt moving soil bound Pb further down
this catchment and bypassing the lakes altogether due to lake ice. Kim et al. (2015) measured dissolved
Pb (and other trace elements) from melting glaciers along the Antarctic coastline in Marian Cove and
determined glacier meltwater is a significant source of Pb to the cove. In ice, Pb was measured at values
between 90-920 pM and 88-550 pM in snow. Pb ranged from 30-120 pM in seawater and Pb
concentration decreased with increasing salinity in seawater samples.
As described in Section 1.2.4, fires are a large and increasing contributor to ambient air Pb
concentrations. A few recent studies have specifically examined Pb (and other metals) movement
following fires. Overall, metal mobility to waterways following a burn event is largely dependent upon
the first storm event following the fire. A review of metal mobilization following fire found fires can lead
to mobilization of Pb into waterways (Abraham et al.. 2017). Pb bound within the soil matrix, in
vegetation, or other burned materials, is released following a burn and is readily washed into downstream
waterbodies with rainfall. Burton et al. (2016) found the total Pb (unfiltered) median water concentrations
downstream of the Station Fire in California was higher than concentrations outside the burn area and
higher than concentrations measured prior to storm events. Furthermore, these post-fire total Pb
concentrations exceeded the recommended aquatic use criteria (CCC) following storm events. The
authors suggest these higher Pb concentrations within burned areas were likely due to ash runoff because
Pb concentration in ash samples was higher than soil samples. Pb source in the ash was primarily
vegetative or biogenic in origin but Pb in ash collected from residential burn areas was higher. A 2011
global review that examined water quality in forested catchments following wildfire reported Pb exceeded
water quality guidelines in Australia and the United States and is associated with high suspended solids
concentrations following post-fire rainfall events (Smith et al.. 2011).
1.3.3.3. Sedimentation, Transport, and Flux in Water and Sediment
As in the 2006 Pb AQCD and the 2013 Pb ISA (U.S. EPA. 2013. 2006). chemical and physical
characteristics of the water such as pH, salinity, and flow rate as well as the chemical and physical
properties of the suspended sediments determine the fate of Pb and therefore influence rates of
sedimentation and flux as well as transport downstream. Recent publications provide additional support
regarding Pb adsorption onto organic-rich or small colloid particles as well as the importance of water
flowrate in settlement and downstream transport. Literature since the previous ISA provides new detail of
the effects of seasonality on Pb fate and transport in water. Seasonal patterns of precipitation can lead to
differences in runoff, flowrate, and turbidity, for example, which can subsequently alter sedimentation
rates, transport downstream, and flux from sediments.
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1.3.3.3.1. Urban
Within urban environments, city infrastructure can lead to increased loading and movement into
downstream reaches. In an urban Baltimore area watershed, sediment concentrations of metals including
Pb increased with urbanization (Bain et al.. 2012). Storm water flow and well-drained soils with low OM
interact to increase runoff and downstream movement of Pb. In an arid California watershed, urban
infrastructure allows for the quick movement of water away from urban areas resulting in increased Pb
concentrations in receiving water bodies following rain events (McKcc and Gilbreath. 2015). Pb
concentration was correlated with water turbidity due to Pb being primarily within the particulate fraction.
Another study examined the influence of runoff and other diffuse pollution sources on lake water and
sediment chemistry of Hough Park Lake in Central Missouri (Ikem and Adisa. 2011). The lake is
surrounded by a golf course with little natural buffering between the course and the lake. Average lake Pb
concentrations in sediments were 11.05 mg Pb/kg for littoral zone sediment and 0.79 mg Pb/kg in pelagic
zone sediment. Total Pb in the water column was 0.0004 mg/L in the spring. Pb content was primarily in
the residual phase (75%) and the authors suggest since all heavy metals were primarily within the residual
phase (lowest mobility phase), the source of heavy metals is likely due to erosion and runoff of parent
rock material.
Seasonal and local weather patterns interact with other factors such as soil and sediment physical
and chemical characteristics to transport Pb into and within water bodies. One study sampling water
column and surface sediment Pb concentrations in Lake Pontchartrain along I-10 shows seasonal
differences (Zhang et al.. 2016b). Spring sediment concentrations ranged from 16.42-28.25 mg Pb/kg and
from 6.94-21.79 mg Pb/kg during the summer. Water column Pb concentrations in spring ranged from
4.65-7.4 |ig Pb/L and from 4.7-10.4 |ig Pb/L during the summer. The higher sediment concentrations in
spring and higher water column concentrations in summer may be due to warmer summer water
temperatures releasing more Pb from sediments. Differences between spring and summer could also be
due to less precipitation and sediment disturbance via turbidity during cooler months. Pb in sediments
was primarily in the stable residual fraction. In the San Juan River delta of Lake Powell, sediment loading
and associated Pb contamination in the downstream reaches reflect an interaction of seasonality and
precipitation (Frederick et al.. 2019). During the spring, high sustained flow from snowmelt occurs at
high elevation tributaries with a history of mining. This increased flow in the spring contributes more Pb-
contaminated sediment to the downstream delta area. Tributaries that contribute greater sediment during
short rainfall events have lower Pb concentrations.
Transport and settlement patterns of Pb are also a function of sediment particle size. In a study
examining the downstream transport of heavy metals from the superfund site Iron Mountain into the
Sacramento River, Pb enters the Keswick Reservoir primarily in the dissolved form and is precipitated or
adsorbed into the particulate phase (Taylor et al.. 2012). However, Pb does not settle out of the water
column and is instead transported far downstream due to particle association with the colloid phase. This
transport of fine contaminated particles occurs during both high and low flow conditions. In the Miami
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River in Florida, Pb was negatively correlated with sediment particle size (Tanscl and Rafiuddin. 2016).
As particle size decreased, Pb content increased with 900 mg Pb/kg found within the fine sediment
fraction. Due to Pb being bound to finer sediments, turbidity from boating, tidal action, and rain events
are of concern for resuspension and mobilization of Pb within the water column. Following a dam
removal on the Pawtuxet River in Rhode Island, fluxes of all metals including Pb increased in response to
river flow (Katz et al.. 2017). As river flow increased, sediments were resuspended into water and this
particle-bound Pb moved downstream into Narragansett Bay. Sebastiao et al. (2017) examined Pb (and
other metals) in the river sediments at two paved river fords in suburban Philadelphia and across seasons.
Pb was found in higher concentrations in April, July, and in December, but these higher values were not
correlated with any rain event. Pb was positively correlated with organic content but only at the ford that
was less used. The authors suggest this relationship occurs due to less water movement from traffic which
in turn allows Pb to adhere to organic particles and settle out of the water column. Furthermore, since
sediment Pb concentration was still high during the winter when the fords are closed to traffic, Pb
persistence in this system is not seasonal.
1.3.3.3.2. Non-Urban
Seasonal effects such as snowmelt can impact Pb movement. Anthropogenic Pb per unit area in
high elevation lake sediments in the Pyrenees was lower than in the surrounding catchments, and this is
potentially due to Pb deposition largely accumulating in snowpack followed by melting and outwash into
lower elevation systems before the ice on the high elevation lakes can melt and incorporate this deposition
portion as happens at lower elevation lakes (Bacardit et al.. 2012). In a study that measured Pb in surface
sediment at the river mouth of the Papaloapan River across a gradient of increasing water depth in the SW
Gulf of Mexico, Pb was positively correlated with organic carbon but only during the winter months
(Rosales-Hoz et al.. 2015). During the summer, Pb was strongly positively correlated with AI2O3 (also
during the winter months though not as strongly). Pb concentration increased with water depth as did the
muddy proportion in sediments. Muddy sediment output from the river was greater during the summer.
As in urban aquatic environments, environmental and physical drivers of water flow patterns
largely govern the transport of Pb within non-urban aquatic environments with additional influence of
sediment and water chemistry. Pb water concentration at European bog and peaty riparian sites was
positively correlated with DOC (particularly at the bog location) (Broder and Biester. 2017). However, at
this location, Pb concentrations in water were lowest during the spring snowmelt likely due to a dilution
effect whereas during high rainfall flow events, amounts exported increase while water concentrations
decrease. Pb was more likely mobilized due to OM decomposition than affinity for forming Pb-OM
complexes since all elements showed similar patterns regardless of OM affinity. Pb concentrations in
water were highest during the fall when dry periods were followed by high rain events. High elevation
peat mires, soils, and down-catchment reservoirs were sampled in the Snowy Mountains in southeast
Australia (Stromsoe et al.. 2015). Pb input to peat mires is dominated by atmospheric deposition and
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showed greater enrichment of Pb compared with down-catchment reservoirs. Down-catchment reservoirs
had depleted Pb levels in comparison due to a dilution effect of soil-bound Pb and large catchment areas.
In an arctic peatland, Pb aqueous concentrations were 2-3 times higher in the spring compared
with the summer (Stolpc et al.. 2013). Pb concentrations were correlated with DOM. Pb was also
primarily in the 0.5-4 nm colloid fraction. During the spring melt in an artic peatland, pH and high
dissolved DOC occur due to erosion of acidic OM and fine particles while concurrently diluting the
contribution of the bicarbonate parental material to waterways. During the summer, alkalinity increases
while DOC decreases because water inputs shift to groundwater source. A European study examined trace
metal budgets across 14 forested catchments over the period of 1997-2011 (Bringmark et al.. 2013). Due
to high anthropogenic deposition for decades, Pb accumulation in catchment soils was high. Pb is bound
to soil OM in these soils leading to high retention of Pb in the system. At higher altitude sites which
experience greater precipitation, retention is lower due to greater runoff and transport of PM out of the
system. In a similar study in England, while Pb deposition has decreased in recent decades, legacy Pb in
peat catchments is a continuing source of Pb to waterways (Rothw ell et al.. 2011). Atmospheric
deposition was measured at 34 gPb/ha/year while fluvial outputs were 316 gPb/ha/year. Following storm
events, Pb runoff into waters occurs primarily as suspended particles (261 gPb/ha/year) with a smaller
aqueous portion (55 gPb/ha/year).
1.3.3.4. Temporal Trends Documented in Sediments
Temporal trends of Pb deposition in sediment show distinct leaded gasoline peaks in the United
States. These peaks are found globally, corresponding to the specific phase-out periods for multiple
countries. Patterns of increasing Pb concentration occurring from the mid-19th century through the
mid-20th century due to early industry as well as agriculture, weathering, and mining operations are
identifiable in North American lake and reservoir sediments. Following the peak deposition period in the
1960s due to leaded gasoline in North America, widespread decreases in Pb concentration in sediments
are seen over the following half century, but concentration values are still higher than background levels
showing continued deposition, non-point contamination, and/or legacy Pb runoff contributions.
Sediment dating of a mill pond in eastern Virginia shows local Pb sources (weathering and coal
combustion) were the primary inputs to Lake Matoaka during the years 1700-1775 (Balascio et al.. 2019).
In 1780, Pb accumulation decreased slightly, possibly due to a decline of industry which coincides with
the capital of Virginia moving from nearby Williamsburg to Richmond. Over the following two centuries,
Pb accumulation increased, and sources were from regional mining and Pb ore smelting activities. Pb
concentrations continue to increase during the 1900s to a peak maximum in 1975 followed by a sharp
decline. This rise and fall of Pb accumulation reflect the increase in coal combustion, smelting, and use of
leaded gasoline. Sediment records in Lake Anna in Virginia have higher Pb concentrations in the
downstream portion of the reservoir (often exceeding 50 mg Pb/kg) in comparison to the upper reaches
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(Odhiambo et al.. 2013). Furthermore, these higher concentrations in the downstream portions were
limited to the younger surface sediments. Lake Anna sits in a rural watershed and sediment cores do not
show the typical increase followed by sharp decrease indicative of leaded gasoline deposition. Instead,
sediment enrichment of Pb in addition with cadmium (Cd), Cu, and Zn point to mining runoff as the
source of Pb enrichment in sediments. An old sulfur (S) mine operated in the area until 1877, and a pyrite,
Cu, and Fe mine until 1920. All other mines ceased operation in the 1990s. A study that measured heavy
metal, polychlorinated biphenyl, and polycyclic aromatic hydrocarbon concentrations in dated sediments
of the lower Anacostia River in Washington D.C. found Pb concentration had two peaks: the first
occurred at corresponding depth of 1943 and the second in 1984 (Vclinskv et al.. 2011). The cause for the
early 20th century peak is not clear, as Pb sources in the area included both agriculture and industry. The
second peak in 1984 and subsequent decrease likely corresponds to the use and then phase-out of leaded
gasoline. A location near the Navy Yard and Government Services Administration showed an increase in
Pb concentration again over the years of 1989-2000, and the authors note this sampling location is close
to large storm water and sewer drainages.
In Horseshoe Lake near St. Louis, three main periods of variable Pb pollution were identified.
The first period was dated as pre-settlement, had low Pb concentration and the lowest 2u6Pb/2u7Pb ratio and
is representative of background parental material and deposition from flooding of the Mississippi River
basin (Brugam et al.. 2012). The second period dated as post 1750 had increasing Pb concentration, and
the 2u6Pb/2u7Pb ratio diverges from the ratio in the first period. The 2u6Pb/2u7Pb ratio increases and matches
Missouri ore samples, and this period coincides with the start of Pb mining in the region. The third period
dated from 1915 to the present also contained high Pb concentrations but a lower 2u6Pb/2u7Pb ratio than
period two. The source of Pb in this sediment layer is less clear because the 2u6Pb/2"7Pb ratio is similar to
vehicle exhaust values of leaded gasoline but is also similar to a nearby old Pb smelter. The lower ratio
but high concentrations may also be reflecting erosion and settling of upstream agricultural runoff. The
parental material upstream under agriculture has a lower ratio than silt from the Mississippi River but a
higher Pb concentration due to agriculture. Lake Whittington, an oxbow lake, was created in 1937 by the
U.S. Army Corps of Engineers off the Mississippi River. Flooding of the Mississippi River is the main
source of sediment to Lake Whittington (Van Metre and Horowitz. 2013). Sediment analysis within the
lake shows Pb concentration increased from 1938 to the 1970s followed by a decrease. The increase is
due, in part, by greater sediment contribution from polluted up-river watersheds of the upper Mississippi
and Ohio rivers and less contribution from the cleaner Missouri River watershed which had extensive
damming in the 1950s. The concentration decrease post 1970s is explained by reduction in leaded
gasoline emissions.
Sediments were sampled and dated in an oxbow lake in southwestern Pennsylvania to establish
historical contamination in an area of the country with a long pollution exposure history (Ostrofskv and
Schworm. 2011). Pb concentrations increased from 1915 to 1938 corresponding to the opening Donora
Zinc Works. Pb levels decrease during the 1940s corresponding depth layer which the authors suggest
reflects either a decrease in production or improvement in recovery methods. This downward trend
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continues through the 1950s (Donora Zinc Works closed in 1957), but Pb concentration increases shortly
thereafter around the time a coal powerplant opened nearby. Pb concentration decreases again in the
1980s—perhaps in response to the cessation of leaded gasoline. However, As, Cd, and Zn concentrations
also decrease suggesting the Pb pollution patterns in this area during this time are instead linked to the
coal powerplant (Rossi et al.. 2017). In Sandy Lake, Pennsylvania, Pb levels increase alongside Fe, Mn,
and S from approximately 1770 until the 2000s. The increase in concentrations seen in Pb and other
elements corresponds with the opening of a coal mine which directly contaminated Sandy Lake with acid
mine drainage in the late 1800s. The decrease in Pb levels in the 2000s likely reflects the decrease in
deposition from leaded gasoline. In another study, Pb sediment concentrations in Little Lake Bonnet in
Florida increased over the period of 1874 to 1920 with a peak of -28 mg Pb/L (Escobar et al.. 2013).
Concentrations increased further to ~38 mg Pb/L in 1949 with an overall peak in 1990 of -12 mg Pb/L.
Pb concentration then declined after 2001 to -60 mg Pb/L. Little Lake Jackson in Florida showed similar
patterns with a peak between mid-1970 and early 1980s. Isotope ratio analysis ties Pb peak patterns with
leaded gasoline. The general increasing concentrations in the lakes during the 20th century correspond
with broad regional industrialization during this period including pesticide and fertilizer use on golf
courses, Pb-As insecticide use on nearby citrus groves, and proximity of coal power plants.
A detailed sediment analysis from Vermillion Lake in Sudbury, Ontario, Canada linked Pb
concentration to historical industry and leaded gasoline (Schindler and Kamber. 2013; Wiklund et al..
2012). Pb concentrations first started increasing in the late 19th century at a time when logging in the area
first started. Pb concentrations increased until apeak in the late 1960s into the 1970s. The subsequent
decrease in concentration corresponds to the phase-out of leaded gasoline in Canada in 1976 and union
strikes at nearby mines, resulting in low production. Pb ratios indicated multiple sources of Pb. The parent
rock and ores have a unique 2u6Pb/2"7Pb ratio and patterns in this ratio combined with increasing nickel
(Ni) concentrations over the period of 1905-1919 indicating Pb level was primarily due to mining
contamination. The continuing increase in Pb concentrations after this period would likely reflect greater
deposition from nearby smelters and refineries which has a similar ratio profile as leaded gasoline. Child
et al. (2018) examined sources and geographic extent of atmospheric metal deposition across eastern
Washington lakes within 50 km from the Trail smelter in British Columbia. Pb isotopes and deposition
profiles indicate the Trail smelter as a primary source of atmospheric trace metal deposition including a
lake outside the 50 km radius and upwind (Louchouarn et al.. 2012). A study by Dunnington et al. (2020)
examined Pb trends using sediment dating in multiple lakes across northeastern North America. Sampled
lakes included locations in the Adirondacks (northeast New York), lakes across Vermont, New
Hampshire, and Maine, as well as several lakes in Nova Scotia. In general, Pb concentrations decreased
from west to east. Sediment dating reveals anthropogenic Pb concentrations began increasing first in the
Adirondack region in 1859 followed by the VT-NH-ME region in 1874, and finally in Nova Scotia in
1901. Authors acknowledge early Pb emissions were likely due to coal combustion in the Adirondack and
New England lakes, but with the increase in Nova Scotia lakes after 1923, Pb from gasoline was likely an
important deposition source to all lakes in this study. Furthermore, looking across the whole sediment
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core, Pb in youngest sediment is still higher than pre-industrial levels—an indication of continued
contamination and loading from legacy Pb in runoff.
In a review by Marx et al. (2016). global contamination records of Pb were examined using
sediment, peat, and ice cores from across North and South America, Europe, Asia, Australia, and both
polar regions. In North America, Pb contamination dating back to 6500 BCE was found and the authors
link this early contamination to pre-historic Cu use. Enrichment in this core increases between 1300 CE
and 1500 CE, where enrichment doubles, corresponding to the start of the industrial revolution. By the
1960s CE, enrichment peaks are followed by a decline to 2002 (though enrichment is still well above the
1500 CE values). A peat and lake core in Canada record enrichment starting much later (1800s CE) and
linking it to coal mining. The peat core shows Pb enrichment peak in 1910 while the lake core peaks in
the 1970s. Overall, Europe and North America have higher enrichment values than Australia and
Antarctica. Globally, Pb enrichment starts in pre-historic era in Europe, North America, and east Asia
while in South America, enrichment starts during the Middle Ages. However, by the latter half of the
1800s, Pb enrichment is globally well above background levels, increasing until the 1970s when
enrichment declines in Europe and North America but continues to increase in east Asia and Australia. A
global study by Zhou et al. (2020a) sampled 168 rivers and 71 lakes from 1972 to 2017 to examine global
patterns in heavy metal pollution. Pb levels globally increased during the 1970s and 1980s to a peak level
of 257 |ig/L in the 1990s. Pb levels decreased slightly during the 2000s with a slight increase in the
2010s. Pb levels were highest in South American water bodies (-333 |ig Pb/L) and lowest in Europe
(-14 (ig Pb/L). Potential sources included rock weathering, fertilizer and pesticide application, mining
and manufacturing, and waste discharge. Across the decades, the primary source of Pb was determined to
be from mining and manufacturing.
1.3.3.5. Sediment Pb Pools as Potential Sources to Surface Waters
The removal or breeching of worn-down dams, such as old mill dams, in the eastern United States
are a potential new source of legacy Pb for downstream waterways. One study by (Niemitz et al.. 2013)
found higher Pb sediment levels above a former mill dam draining former and current agricultural lands
compared with dammed sediment above a forested catchment within the upper reach of the Yellow
Breeches Creek watershed in Pennsylvania. Sediments at both locations, however, contained legacy Pb
from gasoline emissions. In another study, while Pb flux initially increased following the removal of
Pawtuxet River dam and was positively correlated with river flow, SPM concentration decreased and
remained lower than pre-removal levels for the duration of the study (Katz et al.. 2017). This decrease in
SPM concentration, however, is potentially due to increased contribution of low-contamination level soils
previously above the flood zone downstream of the dam site, and sediment cores taken in multiple
locations upstream exceeded sediment quality criteria. Therefore, while removal of a dam may lead to a
dilution effect on downstream waters, pools of legacy Pb upstream of a dam will still move into
downstream waterways with the post-removal associated increase in water flow. Furthermore, these
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contaminated sediments can transform to more biologically available forms as they contribute to
increased turbidity with increased water flow and salinity as with the Yellow Breeches and Pawtuxet
River's flow to the Atlantic (Chesapeake and Narraganset Bay, respectively). Interestingly, in a study
examining the effect of beaver dams on heavy metal retention and sediment contamination found Pb
aqueous concentration to be lower in the outflow because the dammed water acts as an oxidation pond
and results in sorption of Pb to iron oxides (Shepherd and Nairn. 2020). In another study that examined
trace metals in surface sediments of a tidal tributary of the Chesapeake Bay, Pb concentrations were
above threshold effects levels at 44% of sites sampled. The Chester River watershed is a forested and
agricultural watershed and metal accumulation within riverine sediments is likely due to a combination of
multiple non-point source runoff as well as tidal exchange with the Chesapeake—highlighting metal
pollution transport can occur upstream in estuarine environments (Krahforst et al.. 2022). Hurricane
Sandy associated storm surge resuspended Barnegat Bay sediments rich in metal contaminants and
transported them northward within the Bay (Romanok et al.. 2016). The source of metal contaminants
within sediment silt comes from runoff from inundated urban lands, wetlands, estuaries, streams, and
from the resuspension into water of estuarine sediments that may have previously been considered pools
of legacy contaminants—including Pb from upstream deposits. Similarly, as discussed in
Section 1.3.3.2.1, runoff and sediment concentrations have been linked to characteristics of urbanization
suggesting urban areas are a potential source of resuspended Pb in water from sediments following storm
or high-flow events.
1.3.3.6. Summary of Pb Fate and Transport in Aquatic Systems
In summary, literature since the 2013 Pb ISA supports previous conclusions regarding the
physicochemical drivers of Pb fate and transport in aquatic systems. Studies continue to report runoff
from urban or historically industrial areas contain higher Pb concentrations than non-urban areas with
new information highlighting relationships between street length and density, population density, and
land cover with runoff. Recent studies expand on the influence of seasonality and precipitation events on
runoff as well as transport and sedimentation. Timing of snow and ice melt can alter down-catchment
transport of Pb in high elevation watersheds, for example, while another study found water column
concentrations differed between summer and winter—possibly due to differences in precipitation patterns
influencing sedimentation and resuspension into water. A collection of recent studies linked Pb
concentration peaks in lacustrine and riverine sediment cores to national and global patterns of
industrialization in the late 19th and early 20th century, to increased vehicle abundance and associated
leaded gasoline in the mid-20th century, followed by a decline in Pb concentration coinciding with the
phase-out of leaded gasoline and stricter emissions regulations. Furthermore, new literature also
addressed the importance of turbidity and resuspension into water in relation to legacy Pb pools. While Pb
deposition has decreased in the last half century with the phase-out of leaded gasoline and stricter
regulation, Pb sediment pools in areas with a history of industry and urbanization are vulnerable to
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resuspension into water and both down and upstream movement following a disturbance event. Dam
removal or other disturbances to streams in the eastern United States can lead to resuspension in water
and dissolution of Pb-contaminated sediment that was previously deposited. Lastly, with the predicted
increase in drought alongside less frequent but more severe precipitation patterns across most of the
United States, the potential for remobilization of legacy Pb is a growing area of concern and
consideration.
1.3.4. Fate and Transport in Urban Media
Additional media besides air, water, and soil are useful for understanding how Pb moves and
changes overtime in the urban environment. These can include urban soil (Section 1.2.7), resuspended
airborne dust, road dust (Section 1.2.5), and house dust, between which Pb can be transported or cycled.
Pb concentrations are characteristically higher in urban soil than other soils. Frank et al. (2019) reported
the mean estimate of Pb concentration in urban residential soils was 3 times greater than for non-urban
soils in a meta-analysis of recent studies. Obeng-Gvasi et al. (2021) estimated that urban soil Pb
concentrations were approximately 7 times U.S. background levels (Section 11.1.3) through a comparison
of an analysis 84 studies of U.S. soil Pb with a U.S. Geological Survey study of U.S. background soil Pb
concentrations (Datko-Williams et al.. 2014; Smith et al.. 2013). Datko-Williams et al. (2014) concluded
that there had been little change in urban U.S. soil Pb concentrations from 1970 to 2012. The highest
concentrations often occur in roadside soil and near buildings, reflecting the proximity to legacy Pb from
leaded gasoline and deteriorating paint, respectively (Obeng-Gvasi et al.. 2021). although Frank et al.
(2019) reported lower soil Pb concentrations for roadside soils than for residential urban soils in their
meta-analysis of studies on soil Pb concentrations in the United States. The 2013 Pb ISA also reviewed
observations of decreasing soil Pb concentrations with distance from a road (U.S. EPA. 2013). There is
some evidence that the pattern of decreasing soil Pb concentrations with road distance are paralleled by
near-road air Pb concentrations (Section 1.2.6).
Pb has a very long residence time in soil, with models predicting more than a century until soil
concentrations return to steady-state levels (Harris and Davidson. 2005). This is consistent with the slow
transport rates typically observed for a range of conditions (Section 1.3.2) and facilitates the persistence
of long-term hot spots. For example, as described in Section 1.2.7, soil Pb in Durham NC ranged from 6-
8825 mg Pb/kg and soil Pb concentrations were higher around older homes, indicating contribution from
leaded paint (Wade et al.. 2021). Home age accounted for 40% of the variance in foundation soil Pb, with
soil near painted houses containing significantly higher soil Pb than for brick homes (Wade et al.. 2021).
Studies in East Chicago IL, Greensboro NC, Brooklyn NY, and Philadelphia PA have explored the high
spatial variability of urban soil Pb concentrations, with hot spots related to income and racial disparities
(Caballero-Gomez et al.. 2022; Pavilonis et al.. 2022; Hague et al.. 2021; Obeng-Gvasi et al.. 2021).
Urban and neighborhood-scale spatial variability of ambient air Pb concentrations have been observed in
recent studies, but not directly related to urban soil (Section 1.5.2).
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Resuspension of Pb in contaminated soil and road dust by traffic, construction, and wind has been
described as a potential contributor to airborne Pb under some circumstances (U.S. EPA. 2013). The 2013
Pb ISA reported that contaminants associated with particles with diameters up to about 100 (mi can
typically become resuspended into air, but particles larger than 10-20 (mi typically do not remain
airborne long enough for appreciable transport (U.S. EPA. 2013; Nicholson. 1988; Gillette et al.. 1974).
However, the extent of resuspension of contaminants in surface soil and dust particles into air depends
strongly on landscape, geology, particle size and wind speed. The 2013 Pb ISA summarized factors
influencing resuspension into air in a complex urban landscape with heavy traffic, buildings, pavement
and above- and below-ground infrastructure (U.S. EPA. 2013). The critical diameter at which
resuspension into air occurs is the diameter at which the particle settling velocity becomes equal to the
friction velocity of air needed to move the particle at rest. Although this was estimated at roughly 20 |im
in an open landscape (U.S. EPA. 2013; Gillette et al.. 1974). a higher friction velocity is expected for
urban environments with traffic-induced turbulence (Britter and Hanna. 2003). This could result in
resuspension of somewhat larger particles into air in an urban setting with heavy traffic (Nicholson and
Branson. 1990). Near-road observations indicate that the fraction of total particulate Pb associated with
particles larger than 10 (mi can be 15% or more (U.S. EPA. 2013).
Pb particles from soil and dust occur at sizes ranging from 0.1-10 (mi, a size range with potential
for resuspension into air and inhalation (O'Shea et al.. 2021). Laboratory studies and sampling in areas
with previous major emissions sources suggests the potential for resuspension of near-surface soil-bound
Pb to contribute to airborne concentrations in those areas (Pingitore et al.. 2009; U.S. EPA. 2006). A
previously reported modeling study estimated up to 90% of Pb emissions in Southern California are
attributed to Pb resuspension from soil and road dust into air (Harris and Davidson. 2005). though Lankev
et al. (1998) noted a smaller estimate of 43%. With estimated annual emissions of 54,000 kg airborne Pb
in 2001, modeled resuspension was identified as the largest source of airborne Pb in the Southern
California region (Harris and Davidson. 2005). Although Pb from contemporary sources can also be
resuspended into air, emissions from historical sources were considerably greater, and substantial Pb from
sources like leaded gasoline and historical industrial emissions or from non-atmospheric sources like
paint have accumulated in soil over many years, particularly in urban areas (Section 1.2.7).
Correlations between soil Pb concentrations and atmospheric Pb concentrations have been
observed in recent studies. Laidlaw et al. (2014) reported atmospheric Pb loadings increased by
0.066 |ig/m2/28 days for every mg Pb/kg increase in soil Pb. Another study found that a 1% increase in
estimated resuspended soil concentration in air corresponded to a 0.39% increase in atmospheric Pb
concentration (95% CI, 0.28 to 0.5%) (Zahran et al.. 2013). The isotopic composition of Pb in airborne
particles is consistent with that of road dust and topsoil, with significant contributions (a binary mixing
model found from 32 ± 10 to 43 ± 9%) of Pb from leaded gasoline (Resongles et al.. 2021). Atmospheric
soil and Pb aerosols are 3.15 and 3.12 times higher, respectively, during weekdays than weekends,
suggesting traffic as a major driver of Pb resuspension into air (Laidlaw et al.. 2012). In London, 450-
650 kg/year of Pb is emitted as resuspended dust, a similar magnitude as primary Pb air emissions in
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urban locations (Resongles et al.. 2021). Additionally, Pb isotopic composition was similar for particles
collected at ground-level and building height, suggesting Pb is well mixed throughout the vertical column
in urban environments (Resongles et al.. 2021). Dust emissions are significant and represent missing
sources in the emission inventories (Xu et al.. 2019). To reduce national-scale bias of modeled Pb
concentrations, a fivefold increase in anthropogenic emissions of Pb was necessary to achieve agreement
between simulated and observed ambient air Pb concentrations (Xu et al.. 2019). While these studies
suggest that resuspension Pb from soil into air is a potentially important local source of Pb in ambient air,
it appears to be a much smaller contributor to current ambient air Pb concentrations than leaded gasoline
exhaust was in previous decades. Airborne Pb monitoring was originally required for urban National Core
multipollutant monitoring network (NCore) multipollutant monitoring network sites (Section 1.4) but was
discontinued in 2016 because concentrations were consistently much lower than NAAQS levels (40 Code
of Federal Regulations (CFR) Part 58, Appendix A). Moreover, at five Pb monitoring sites near roads
with heavy traffic, ambient air Pb concentrations decreased from more than 1 (ig/m3 in 1979 to less than
0.03 ng/m3 in 2010 (U.S. EPA. 2013).
Several studies have reported seasonal patterns of resuspension from soil into air, with highest
resuspension occurring in summer and autumn when soils are driest (Resongles et al.. 2021; Mielke et al..
2019; Laidlaw et al.. 2017; Laidlaw et al.. 2016; Laidlaw et al.. 2014). In Detroit, atmospheric Pb is
44.8% higher in August than January (Figure 1-5) (Zahran et al.. 2013). This seasonal pattern is also
observed in measurements of children's blood levels (Laidlaw et al.. 2017; Mielke et al.. 2017; Laidlaw et
al.. 2016; Zahran et al.. 2013) and is discussed in detail in Section 2.4. National scale modeling of heavy
metal concentrations with the chemical transport model GEOS-Chem indicated that simulated heavy
metal concentrations in PM2 5 over continental North America were consistent with monitoring
observations in winter, but generally low in other seasons, suggesting that contributions of missing
sources of Pb follows a seasonal pattern similar to that observed for airborne soil components (Xu et al..
2019).
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Air Pb
Air Pb (Median Spline) ¦ Air Soil
Air Soil (Median Spline)
Source: Reprinted with permission from Zahran et al. (2013). Copyright 2013, American Chemical Society.
Air soil refers to the estimated ambient air concentration of soil-derived PM based on crustal element concentrations. Weather-
adjusted concentrations are concentrations that have been adjusted for relative humidity, pressure, temperature, visibility, and wind
speed using their known relationships with air Pb and air soil to determine their seasonality independent of short-term weather
conditions. The median spline is a smoothing function based on a polynomial fit.
Figure 1-5 Ambient air Pb and air soil concentrations and median splines in
(jg/m3 from Detroit, Michigan.
Though rare, extreme weather events can alter soil Pb concentrations drastically. Following
Hurricane Katrina in New Orleans, soil Pb decreased from 285 to 55 mg Pb/kg on public land and from
710 to 291 mg Pb/kg on private land (Mielke et al.. 2017). These observed effects are likely due to result
in decreased resuspension into air following flooding, as well as transport of soil from outside the city
covering Pb-contaminated urban soil (Mielke et al.. 2017). Recent Pb isotopic aerosol signatures show
origins from leaded gasoline, suggesting Pb sources have not changed substantially since the removal of
leaded gasoline (Resongles et al.. 2021). Pb in PMio samples collected in London ranges from 3.9-
19.4 ng/m3, with deposition rates of 11,700-45,800 ng/m2/day for Pb associated with total suspended
particulate (TSP) matter (Resongles et al.. 2021). Thirty-one percent of PMio particles measured were
attributed to resuspended road dust in air (Resongles et al.. 2021).
The larger size of resuspended dust particles compared with typical atmospheric particle size
distributions makes the atmospheric lifetimes and travel distances of the airborne dust Pb potentially
shorter than those expected for PM2.5 or PMio (U.S. EPA. 2013). As a result, resuspension of large
amounts of soil Pb into air does not appear to be an efficient process for Pb removal from a neighborhood.
However, resuspension followed by relatively rapid deposition provides a potential process for Pb to
translocate within neighborhoods, reducing high concentrations near busy roads while increasing it in
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other areas. This pattern of evening out soil Pb concentrations in city centers over long time scales has
been described by Laidlaw and Filippelli (2008).
Association with airborne dust also provides Pb with a transport pathway indoors, where it
deposits as house dust. Along with urban soil, house dust has been a particular concern for accumulation
of Pb from deteriorating paint (Lanphear et al.. 1998). The evidence for the link between atmospheric Pb
and house dust Pb near large industrial sources can be strong enough that urban-scale house dust Pb
concentrations have been used to effectively track changes in atmospheric deposition patterns caused by
the addition of a tall stack to a smelter (Van Pelt et al.. 2020). Since there are also other processes for
transport of Pb between soil and house dust, an unknown portion of the Pb in house dust becomes
airborne after its release into the environment. However, there is potential for Pb resuspension into air to
serve as a source of both ambient air Pb and house dust Pb. In a recent study of childhood leukemia risk
from carpet dust, mean dust loadings were 24.5 and 15.3 (ig/ft2 in homes of children with and without
leukemia, respectively (Whitehead et al.. 2015). These results compare to Frank et al. (2019). who
reported a mean of 13 (ig/ft2 for 535 floor samples and 214 (ig/ft2 for 380 windowsill samples in a meta-
analysis of studies published between 1999 and 2015. Gillings et al. (2022) observed that Pb in both
house dust and surface soil decreased with distance from mining areas. The decline with distance was
steeper for soil than for house dust (Gillings et al.. 2022). Whether soil Pb was removed or buried deeper
under the surface was not discussed. There is also evidence for Pb-contaminated dust deposition onto
roofs, which could then undergo resuspension into the air during demolition (Caballero-Gomez et al..
2022).
Recent research supports prior information on the influence of legacy Pb from leaded gasoline,
past industrial emissions, and deteriorating paint on soil Pb concentrations in some areas, with the highest
concentrations near roads and buildings. Within urban soil there appears to be gradual Pb transport away
from roads and buildings, and a slow reduction of soil Pb concentration gradients over time, potentially
due at least in part to cycles of resuspension into air, atmospheric transport, and deposition. There is also
transport between soil and other compartments, namely house dust, road dust, and air. Overall, transport
rates are extremely slow, and changes in concentration gradients near sources and hot spots are very
gradual, apparently remaining for decades without intervention or other external disturbances like
hurricanes or floods. Current research is taking place toward understanding rates of Pb transport and their
influencing factors either among compartments or among different locations in the same medium
(i.e., soil, road dust). One salient observation is that the extremely slow movement of Pb through the
urban environment facilitates persistent hot spots of high soil Pb and house dust Pb concentrations. There
is little data on whether hot spots of high soil Pb and dust Pb concentrations also lead to pockets of high
ambient air concentrations, with the exception of near-road observations. Although not all urban Pb
transport involves air, there is evidence that resuspension of Pb in urban soil may contribute to airborne
Pb concentrations in some locations.
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1.4
Monitoring of Pb in Ambient Air
This section describes advances in development and evaluation of sampling and analytical
methods for monitoring and measurement of airborne Pb. Section 1.4.1 describes national monitoring
networks currently in operation. Section 1.4.2 describes recent research to evaluate the performance of the
Federal Reference Method (FRM) for Pb in ambient air. Section 1.4.3 provides a summary of network
monitoring challenges related to Pb airborne Pb sampling. Section 1.4.4 describes recent advances in
sampling and analysis of airborne Pb in monitoring and research.
1.4.1. Network Monitoring
Four national monitoring networks collect data on Pb concentrations in ambient air and report it
to the Air Quality System (AQS). Up-to-date network descriptions and monitor locations for the Pb state
and local air monitoring stations (SLAMS), Chemical Speciation Network (CSN), Interagency
Monitoring of Protected Visual Environments (IMPROVE), National Air Toxics Trends Stations
(NATTS), and NCore networks are available in "Overview of Lead (Pb) Air Quality in the United States"
(U.S. EPA. 2023c). Measurements between networks are not directly comparable because of PM size
range differences and other differences in sampling and analytical methods.
The SLAMS network shown in Figure 1-6 is operated by state and local agencies and monitors
are sited in compliance with regulatory requirements (U.S. EPA. 2023c). Although data are used for other
scientific purposes, the SLAMS network is designed primarily with the goal of evaluating compliance
with the Pb NAAQS. As described in the 2013 Pb ISA, new monitoring requirements that were revised
over the period from 2008 to 2010 expanded the number of SLAMS monitors (U.S. EPA. 2013). Source-
oriented monitors are required near sources of Pb air emissions which are expected to or have been shown
to contribute to ambient air Pb concentrations exceeding the NAAQS. At a minimum there must be one
source-oriented site located to measure the maximum Pb concentration in ambient air resulting from each
non-airport Pb source estimated to emit 0.50 tons of Pb per year or more and from each airport estimated
emit 1.0 tons of Pb per year or more. These thresholds were established to ensure monitoring takes place
near Pb air sources with the greatest potential to cause ambient air concentrations to exceed the Pb
NAAQS.
In addition to the SLAMS network, Pb is monitored in several monitoring networks designed for
non-regulatory purposes described in Figure 1-7. Pb is also measured within the CSN, the IMPROVE
network, and the NATTS network. These networks are designed to meet other objectives besides NAAQS
attainment evaluation. Pb in PM2 5 is monitored in the CSN and IMPROVE networks. The purpose of the
CSN is to monitor PM2 5 species to help understand spatial and temporal variations in PM2 5 chemistry,
including annual, seasonal, and sub-seasonal trends. Pb is one of 33 elements collected on Teflon filters
every third day and analyzed by energy-dispersive X-ray fluorescence spectrometry (ED-XRF).
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IMPROVE network monitors are operated mostly in rural locations by the National Park Service and
other agencies for establishing current visibility conditions, tracking changes in visibility, and
determining causal mechanisms of visibility impairment in national parks and wilderness areas. As for
CSN samples, Pb in PM2.5 sampled in the IMPROVE network is analyzed by ED-XRF.
As shown in Figure 1-7, Pb is also monitored in the NATTS network, designed to monitor
concentrations of hazardous air pollutants (HAPs). The NATTS is intended to provide model input, to
observe long-term trends in HAP concentrations, and to examine emission control strategies. Pb is one of
seven core inorganic HAPs collected in PM10. on Teflon filters and typically analyzed by IC-PMS. Up-to-
date maps of monitor locations for the Pb SLAMS, CSN, IMPROVE, NCore, and NATTS networks are
available at https://www.epa.gov/air-qualitv-analvsis/lead-naaqs-review-analvses-and-data-sets.
Although not required, some monitoring agencies also conduct non-source-oriented Pb
monitoring at National Core multipollutant monitoring network (NCore) sites. There are 60 such urban
NCore sites in Core Based Statistical Areas (CBSA) with a population of 500,000 or more. For these non-
source-oriented monitors, the main objective is to gather information on neighborhood scale Pb
concentrations that are typical of urban areas to better understand ambient air-related exposures of
populations in these areas. The NCore network is a multi-pollutant monitoring network with advanced
measurement systems for particles, pollutant gases, and meteorology. It is designed to support timely
reporting of data to the public, development of emission strategies, and long-term health assessments.
Monitoring network maps shown in Figure 1-6 and Figure 1-7 are annually updated at
https://www.epa.gov/air-qualitv-analvsis/lead-naaqs-review-analvses-and-data-sets. Monitoring for Pb
was required at some NCore monitoring sites when the network was initiated in 2011.
The number of Pb FRM and Federal Equivalent Method (FEM) monitors has decreased steadily
since 1985, with a brief uptick corresponding to the NCore monitoring requirement for Pb from 2011 to
2016 (see Section 1.4.1, Figure 1-11). The long-term decrease in the number of monitors is due in part to
criteria established to allow Pb monitoring to be discontinued. Requests for discontinuation, subject to the
review of the Regional Administrator, will be approved for any state and local monitoring station monitor
which has shown attainment during the previous five years, that has a probability of less than 10% of
exceeding 80% of the applicable NAAQS during the next three years based on the levels, trends, and
variability observed in the past, and which is not specifically required by an attainment plan or
maintenance plan (40 CFR Part 58.14c). The requirement for monitoring Pb near sources may also be
waived by the Regional Administrator if the state or local agency can demonstrate that the Pb source will
not contribute to a maximum Pb concentration in ambient air in excess of 50% of the NAAQS based on
historical monitoring data, modeling, or other means, and the waiver must be renewed once every 5 years.
Concentrations fell below this level at many source-oriented monitors after the source was eliminated or
emissions reduced. Concentration at non-source-oriented monitors in the U.S. are often extremely low.
For example, the maximum rolling 3-month average Pb concentration over a three-year period never
reached 50% of the NAAQS concentration for any of the 23 NCore monitors for which design values
1-54
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were reported for at least some period between 2011 and 2023, and only 4 NCore Pb monitors ever
reported a rolling 3-month average concentration greater than 0.01 |ig/nr\ barely above the detection limit
of 0.0075 (ig/m3 for the Pb-TSP FRM, for any three-month period (U.S. EPA. 2023b).
• Lead (TSP) LC (130) O Lead (TSP) STP (26) • Lead (PM10) LC (2)
Source: (U.S. EPA. 2023c).
Figure 1-6 Map of U.S. ambient air monitoring sites reporting regulatory Pb
data to U.S. EPA during the 2020-2022 period.
1-55
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• CSN (103) • IMPROVE (149) O NCORE/NATTS (70) • OTHER (40)
Source: (U.S. EPA. 2023c).
Figure 1-7 Map of U.S. ambient air monitoring sites reporting non-regulatory
Pb data to U.S. EPA during the 2020-2022 period.
1.4.2. Federal Reference Methods
In order to be used in regulatory decisions judging attainment of the Pb NAAQS, ambient air Pb
concentration data must be obtained through the FRM, or an FEM defined for this purpose. Accordingly,
for enforcement of the air quality standards set forth under the Clean Air Act, the U.S. EPA has
established provisions in the Code of Federal Regulations under which analytical methods can be
designated as FRM or FEM. FRMs and FEMs for the Pb NAAQS exist for both sample collection and
sample analysis. There are two FRMs for Pb sampling in ambient air: (1) Reference Method for the
Determination of Lead in Suspended Particulate Matter Collected from Ambient Air (40 CFRpart 50
Appendix G), and (2) Reference Method for the Determination of Lead in Particulate Matter as PMio
Collected from Ambient Air (40 CFR part 50, Appendix G). The Pb-TSP FRM sample collection method
measures Pb-associated TSP and is required for all source-oriented NAAQS monitors, and the FRM for
Pb- PMio is accepted for Pb NAAQS monitoring at non-source-oriented monitors in specified situations
(U.S. EPA. 2013).
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The Pb-TSP FRM sample collection method specifies use of a high-volume TSP sampler that
meets specified design criteria (40 CFR part 50 Appendix B). Ambient air PM is collected on a glass fiber
filter for 24 hours using a high-volume air sampler. Variability in high-volume TSP sampler collection
efficiency associated with effects of wind speed and wind direction for particles larger than 10 |im has
been documented since the sampler was first implemented for TSP and Pb-TSP sampling and was a major
focus of the 2013 Pb ISA (U.S. EPA. 2013).
To provide informative background for addressing sampling issues associated with particles
larger than 10 (mi, the 2013 Pb ISA reviewed historical research on large particle sampling as well as
recently developed low-volume samplers with manufacturer-designated TSP inlets. The 2013 Pb ISA
concluded that a high degree of variability had been observed between different models of manufacturer-
designated TSP samplers, that no alternative to the FRM TSP sampler had yet been identified, and that
there was still a need to assess the feasibility of a revised TSP sampling method for efficient collection
particles larger than 10 (mi (U.S. EPA. 2013).
The low-volume Pb-PMio FRM sample collection method specifies use of a low-volume PMio
sampler that meets specified design criteria (40 CFR part 50, Appendix Q). Ambient air PM is collected
on a polytetrafluoroethylene (PTFE) filter for 24 hours using active sampling at local conditions with a
low-volume PMio sampler and analyzed by XRF. Use of the FRM for Pb-PMio is allowed in certain
instances where the expected Pb concentration does not approach the NAAQS and in the absence of
nearby sources of Pb associated with particles greater than 10 (mi diameter.
In addition to FRMs used in the SLAMS network, other methods are used in the CSN,
IMPROVE and NATTS networks, as well as in field studies unrelated to network applications. The most
relevant of these methods are listed and described in the 2013 Pb ISA (U.S. EPA. 2013). In the CSN and
IMPROVE networks, Pb in PM2.5 is collected on Teflon filters by low-volume samplers and analyzed by
XRF (U.S. EPA. 2013). Other sampling approaches, including cascade impactor sampling for Pb particle
size distributions, saturation samplers, and passive samplers to provide high-density coverage for
evaluation of spatial variability were reviewed in the 2013 Pb ISA (U.S. EPA. 2013). and recent
applications to Pb spatial variability and size distributions are reported in Section 1.5. Other analytical
methods applied to Pb were also reviewed in the 2013 Pb ISA, including inductively coupled plasma
atomic emission spectroscopy, ED-XRF, proton induced X-ray emission spectroscopy, X-ray
photoelectron spectroscopy, Pb speciation methods including X-ray absorption fine structure as well as
gas chromatography and high performance liquid chromatography combined with mass spectrometry,
ICP-MS, Pb isotope ratio analysis, and continuous Pb monitoring methods (U.S. EPA. 2013). Few new
advances related to Pb analysis have been reported for these methods, but recent applications of some are
reported in Section 1.5.
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1.4.3.
Sampling Considerations
When the Pb NAAQS was revised in 2008, revision of the FRM from a method sampling Pb-TSP
to a method sampling PM less than or equal to 10 |im in diameter (Pb-PMio) was also considered because
of poor precision and variable collection efficiencies with wind speed and direction for larger particle
sizes associated with the Pb-TSP FRM. The Pb-TSP FRM was developed in the 1950s for collection of
TSP. It has been the FRM for airborne Pb since 1973, but over the first decade after its designation as the
Pb FRM, at least twelve studies reported variable sampler performance depending on several factors,
particularly particle size, wind speed, sampler orientation with respect to wind direction, and extent of
passive collection (Krug et al.. 2017). Variations in Pb-TSP FRM sampler performance could be due in
part to broad acceptable performance ranges and the lack of a strictly defined performance standard for
evaluating TSP samplers (Krug et al.. 2017).
The Pb-PMio FRM is not as vulnerable to sampling errors associated with the Pb-TSP FRM
because it is based on a strictly defined performance standard. Transition to a Pb-PMio-based NAAQS
was not supported, however, because Pb associated with particles larger than 10 |im in diameter can be an
important contributor to airborne Pb exposure, in part because of potential resuspension of Pb in urban or
industrial soil or road dust into air in some locations (Section 1.3.4). Additionally, Pb-PMio can only be
used as a surrogate for Pb-TSP if the loss of particles larger than 10 (mi in diameter can be compensated
by firmly establishing a quantitative adjustment factor based on concurrent Pb-TSP and Pb-PMio
sampling that exhibits long-term stability overtime (Krug et al.. 2017). On the grounds of limited
comparisons of Pb-TSP and Pb-PMio available at the time, it was judged that more data sets were needed
before either national or site-specific relationships between Pb-TSP and Pb-PMio could be established.
Thus, the Pb-TSP sampling method was retained as the FRM because of the importance of including Pb
associated with particles larger than 10 |im in diameter and the lack of a quantitative adjustment factor.
However, a Pb-PMio FRM was developed and permitted for non-source-oriented monitoring sites, where
rolling 3-month average Pb-TSP concentrations were less than 0.1 (ig/m3 and Pb associated with particles
larger than 10 (mi was not anticipated.
Although the term TSP implies collection of airborne particles of all sizes, there is evidence that
the Pb-TSP sampler could still miss particles with diameters larger than certain upper limits for efficient
sampling. Although a TSP sampler collects more particulate mass than a PMio sampler, a substantial
fraction of airborne PM could be missed by both samplers if the size range of Pb-associated particles
extends beyond both of their practical size limits for efficient sampling. As described in the 2013 Pb ISA
(U.S. EPA. 2013). in previous observations the TSP sampling efficiency of 97% for 5 (mi particles
dropped to 35% for 15 (mi particles under the same conditions (U.S. EPA. 2013; Wedding et al.. 1977).
and the cut-point for 50% sampling efficiency was observed to decrease from 50 |im at a 2 km/hour wind
speed to 22 (mi at 24 km/hour (Rodes and Evans. 1985). This suggests that at least under some
circumstances resuspended Pb at the high end of the Pb particle size distribution relevant for resuspension
into air might be collected with less than 50% efficiency. Insufficient data on size distributions of
1-58
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airborne resuspended Pb beyond the size range efficiently collected by the Pb-TSP sampler makes it
difficult to assess either the fraction of Pb potentially missed by a Pb-TSP sampler or the contribution of
Pb from this size range to total airborne Pb. An additional challenge to measuring Pb over its entire PM
size range is that spatial variability is greater for coarse than for fine particles. As described in the 2019
PM ISA (U.S. EPA. 2019). PM2.5 concentrations on an urban scale are more uniform than for PM10-2.5. By
extension, including a substantial amount of mass for particles larger than PM10 is potentially
representative over a smaller area containing a smaller population.
1.4.4. Recent Advances in Sampling and Analysis
The 2013 Pb ISA described several new developments in sampling and analysis of airborne Pb,
including the addition of new FRMs and FEMs (U.S. EPA. 2013). Advances in the development of new
sampling and analytical methods described here stem in part from 2008 revisions of the Pb NAAQS.
Specifically, the 2008 revisions described conditions under which a Pb-PMio FRM could be used as an
alternative to the original Pb-TSP FRM. The trade-off between missing Pb from Pb-TSP from upper end
sampling bias and uncertainty versus the potential for missing a large fraction of airborne Pb using the
Pb-PMio FRM continues to be a driving force for further research. In addition, the lower NAAQS level
established by the 2008 revisions triggered a need for alternative analytical methods like ICP-MS with
better performance characterization at lower concentrations. Recent publications provide essential
characterization and comparisons of both old and new measurement methods. These include advanced
wind-tunnel studies of the original Pb-TSP FRM (40 CFR Part 50 Appendix B) (Vanderpool et al.. 2018)
and other TSP samplers (Krug et al.. 2017). a detailed field comparison of collocated Pb-PMio versus Pb-
TSP performance (Ward et al.. 2019). use of particle-size-fractionated airborne lead (Meng et al.. 2014).
performance evaluation and interlaboratory comparison of the recently designated ICP-MS FRM (FR 78,
page 40,000) (Harrington et al.. 2014). and development of relevant reference materials in a suitable
concentration range for XRF and ICP-MS analysis of airborne Pb (Yatkin et al.. 2016).
Following a surge of studies immediately after its development, little new work to further
evaluate Pb-TSP FRM sampler performance was carried out until Krug et al. (2017) summarized and
compared results from the previous evaluations of sampler performance conducted under various
conditions, and used up-to-date methods to quantify the effects of environmental and operational factors
affecting sampler performance in a controlled wind-tunnel setting using isokinetic samplers operating
alongside the sampler. Krug et al. (2017) reported that: 1) sampling effectiveness ranged from 42% to
92% based on particle diameter, across orientations and wind speeds, an effectiveness range that is
comparable to those reported in some previous studies and smaller than those reported in others; 2)
sampling effectiveness is a near monotonic decreasing function of aerodynamic particle size, as predicted
by sampling theory; 3) wind speed plays a significant role in sampling effectiveness; and 4) sampling
effectiveness varies with sampler orientation to the wind and the variability increases with wind speed,
but not to the extent observed in previous studies.
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Several low-volume, portable commercial samplers have been widely referred to as TSP samplers
because of their lack of internal size fractionation. Described as samplers with manufacturer-designated
TSP inlets in the 2013 Pb ISA (U.S. EPA. 2013). these devices also have potential for airborne Pb
sampling, but have not been as extensively evaluated for collection efficiency as a function of particle
size as the Pb-TSP FRM. Some early efforts to characterize alternative low-volume TSP samplers were
summarized in the 2013 Pb ISA, and the value of testing them as potential replacements for high-volume
TSP sampling was discussed (U.S. EPA. 2013). This testing has now been completed by Vanderpool et
al. (2018). who conducted a study specifically to evaluate size-selective performance of six low-volume
manufacturer-designated TSP samplers at different wind speeds in a controlled wind-tunnel setting. Like
the Pb-TSP FRM, sampling performance generally decreased with both particle diameter and wind speed
for each of the inlets evaluated, and all sampling inlets exhibited some degree of measurement bias for
larger particles and at higher wind speeds (Vanderpool et al.. 2018). On average over most wind speeds,
most samplers collected 75% to 95% of particulate mass expected for particle diameters ranging up to
30 (mi (Vanderpool et al.. 2018). However, these would still need to be evaluated more extensively before
approval as a FRM or FEM and use in the NAAQS monitoring network.
Other recent research includes a detailed assessment of collocated Pb-TSP and Pb-PMio
monitoring results to expand on the limited available data for understanding the relationship between Pb-
TSP and Pb-PMio concentrations and to assess the suitability of a site-specific adjustment factor. Pb-TSP
and Pb-PMio data were collected every sixth day for more than three years from a monitor adjacent to the
Walkill secondary smelter in New York, which had been recently equipped with a wet electrostatic
precipitator to reduce emissions of larger particles. Data from the two samplers were strongly correlated
with an adjustment factor of 1.49, somewhat lower than previous observations for primary smelters from
a small number of samples (Ward et al.. 2019). Ward et al. (2019) confirmed that implementation of a Pb-
PMio monitor at their source-oriented location would lead to underestimation of the total ambient air Pb
concentration without the application of an adjustment factor relating Pb-PMio to Pb-TSP. They also
suggested that development of a generic adjustment factor for all Pb monitoring locations was probably
not possible because of differences in particulate Pb characteristics between different locations and source
emissions. These results indicate that in addition to non-source-oriented sites where the Pb-PMio FRM is
currently used, there is also potential to obtain high quality data using Pb-PMio samplers at some source-
oriented sites. However, this approach also requires demonstration of a stable relationship between Pb-
PMio and Pb-TSP over time, which has yet to be evaluated.
Revision of the Pb NAAQS in 2008 resulted in a NAAQS level that approached the limit of
quantitation for the Pb analysis FRM based on flame atomic absorption spectroscopy (FAAS).
Improvements in sensitivity, precision, throughput capability, and extraction efficiency since the
development of the original FRM provided additional motivation for development of a new FRM
(Harrington et al.. 2014). To address these changes, a new Pb analysis FRM based on ICP-MS was
introduced in 2013 (40 CFR Part 50 Appendix G). The new FRM includes two extraction methods: hot
block HNO3 extraction and ultrasonic extraction in a HNO3/HCI mixture. The FRM was evaluated in a
1-60
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multi-laboratory study that demonstrated acceptable sample stability after extraction; acceptable
equivalency with the FAAS FRM; acceptable intra- and interlaboratory precision; comparability across
relevant filter media; acceptable accuracy for analysis of botanical, geological, and industrial standard
reference materials; and method detection limits of less than 5% of the 2008 NAAQS levels (Harrington
et al.. 2014). Considering these results, the ICP-MS-based FRM for Pb-TSP is considered more
appropriate for the 2008 NAAQS level than the previous FAAS-based FRM for Pb-TSP because of its
good performance at significantly lower ambient air Pb-TSP concentrations (Harrington et al.. 2014).
Another useful recent advance for ambient air Pb measurement was the development of new
reference materials suitable for XRF analysis, which is used as an FEM for Pb NAAQS monitoring as
well as for quantifying Pb concentrations in the CSN and IMPROVE monitoring networks. The new
reference materials are useful for laboratory audits, federal equivalency method evaluation, calibration,
and quality control. They were generated by aerosol deposition over a range of Pb concentrations on
PTFE filters used for ambient air sampling, and good long-term stability was demonstrated (Yatkin et al..
2016). The new reference materials fill a gap in commercially available reference materials because
previously available reference materials were not similar to filter media used for collection or the PM
matrix and contained Pb amounts that were not similar to typical ambient air Pb samples. Yatkin et al.
(2016) also used the new reference materials to conduct an interlaboratory comparison of XRF analysis
methods and to establish equivalence between XRF and ICP-MS Pb analysis methods.
Several advances have also recently taken place in research instrumentation to improve time
resolution. Performance was evaluated for an updated version of the Semi-continuous Elements in
Aerosol Sampler (SEAS-III). With this sampler, a high volume of sample is collected in a slurry for
analysis by ICP-MS. A collection efficiency of 87 + 16% was reported and collocated precision was
better than 25% for 20 elements. For Pb, the collocated precision was 33% for sample concentrations
averaging 4 ng/m3, or about 5 times the reported minimum detection limit of 0.79 ng/m3 (Pancras and
Landis. 2011). A portable XRF monitor with subdaily time resolution has also been evaluated and applied
to field measurements of airborne PM (Sofowote et al.. 2019; Tremper et al.. 2018; Furger et al.. 2017).
Extractive electrospray ionization combined with Time-of-flight mass spectrometry has also been
developed for real-time measurement of Pb in ambient air (Giannoukos et al.. 2020).
1.5 Ambient Air Pb Concentration Trends
This section summarizes ambient air Pb concentrations and trends. The 3-month average ambient
air Pb concentrations presented here were created using a simplified approach of the procedures detailed
in 40 CFR part 50 Appendix R and, as such, cannot be directly compared with the Pb NAAQS for
determination of compliance. Section 1.5.1 presents nationwide ambient air Pb concentration trends.
Figure 1-11 summarizes results from numerous, mostly local studies on urban and neighborhood spatial
1-61
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scale variability. Sections 1.5.2, 1.5.3, and 1.5.4 report the latest results on seasonal/diurnal trends, size
distributions, and background concentrations of Pb in ambient air PM from diverse locations.
1.5.1. National Scale Ambient Air Concentrations and Long-Term Trends
Figure 1-8 is a national map of maximum rolling 3-month average Pb concentrations in counties
with Pb-TSP monitors during the period 2020-2022 (U.S. EPA. 2023c). Elevated 3-month average Pb
concentrations were observed in several U.S. locations during the period. Maximum rolling 3-month
average Pb concentrations for the period 2020-2022 exceeded 0.15 (ig/m3 for Canton/Stark County OH
(0.40 (ig/m3), Arecibo PR (0.35 (ig/m3), Troy/Pike County AL (0.22 (.ig/rn3). and Hammond/Lake County
IN (0.16 (.ig/rn3). In each of these locations, daily concentrations occasionally reached or exceeded
1 (.ig/rn3. For example, in 2022 the daily Pb concentration reached or exceeded 1 (.ig/rn3 for 1 day in
Arecibo and Troy AL, and for 2 days in Canton OH and Hammond IN. In Iron County MO, no rolling 3-
month average greater than 0.15 (.ig/rn3 was observed during 2020-2022, but in 2022 the daily
concentration exceeded 1 (.ig/m3 on 7 days. The 2020-2022 maximum rolling 3-month average Pb
concentration was less than 0.15 (.ig/rn3 and the daily Pb concentration was less than 1 (.ig/rn3 at all other
monitoring sites in the United States. In cases where the highest concentrations were observed or where
concentrations exceeded 0.15 (.ig/rn3 for the first time, additional actions were taken. For example, in
Canton OH, where the highest maximum rolling 3-month average Pb concentration for 2020-2022 was
observed, as well as in Hammond IN, where a maximum rolling 3-month average Pb concentration
greater than 0.15 (.ig/m3 w as observed for the first time during the 2020-2022 period, local sources were
sent notices of violations in 2018. In Hammond IN, an area near the source was also designated as a
Superfund site in 2023.
A detailed summary of Pb concentrations measured in the U.S. from monitors used for regulatory
compliance (see Section 1.4.1) is provided in Table 1-3. These are compared to Pb concentrations from
non-regulatory monitoring networks (see Section 1.4.1) in Table 1-4. Maps and tables like Figure 1-8 and
Table 1-3 and Table 1-4 for the most recently available 3-year period will be updated annually and made
available at https://www.epa.gov/air-qualitv-analvsis/lead-naaqs-review-analvses-and-data-sets. Although
maximum and 99th percentile concentrations in Table 1-3 and Table 1-4 exceed 0.15 |ig/nr\ median (p50)
concentrations are under .010 (.ig/m3 for all monitors, including both source and non-source-oriented
monitors. These concentrations are low enough to make trends difficult to discern, especially for non-
source-oriented monitors, and many of the Pb monitors installed at NCore monitoring sites (see
Section 1.4) were later removed after consistent observations of Pb concentrations well below NAAQS
levels. As described in Section 1.2.1, ambient air Pb measurements were also required at 17 U.S. airports
with estimated Pb emissions between 0.50 and 1.0 tons Pb per year for a full year ending in December of
2013, in addition to airports already required to meet ambient air Pb monitoring requirements because of
estimated Pb emissions greater 1.0 tons Pb per year. Airport Pb emissions depend on the level of piston-
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aircraft activity and patterns of runway use. Maximum 3-month average Pb concentrations ranged from
0.1 to 0.33 (ig/m3 and exceeded 0.15 (ig/m3 at 2 of the 17 airports (U.S. EPA. 2015).
• 0.06 - 0.10 ug/mA3 (11 sites) O 0.16 - 0.20 ug/mA3 (1 site)
Source: (U.S. EPA. 2023c).
Figure 1-8 Pb maximum rolling 3-month average in //g/m3 for the 2020-2022
period.
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Table 1-3 Distribution of regulatory Pb-total suspended particle
concentrations in pg/m3 for 2020-2022
metric
quarter
N.sites
N.obs
mean
SD
min
pi
p5
plO
p25
p50
p75
p90
p95
p98
p99
max
max.site
daily
all
130
25,155
0.023
0.079
0.000
0.000
0.001
0.001
0.003
0.006
0.016
0.045
0.086
0.178
0.295
2.923
290930021
daily
1st quarter
127
6,276
0.022
0.072
0.000
0.000
0.001
0.001
0.002
0.006
0.016
0.041
0.082
0.174
0.293
2.019
290930021
daily
2nd quarter
129
6,165
0.027
0.100
0.000
0.000
0.001
0.002
0.003
0.007
0.018
0.053
0.100
0.214
0.344
2.923
290930021
daily
3rd quarter
129
6,489
0.022
0.069
0.000
0.000
0.001
0.002
0.003
0.007
0.017
0.045
0.083
0.160
0.265
2.370
391510024
daily
4th quarter
128
6,225
0.022
0.073
0.000
0.000
0.001
o.ooi
0.002
0.006
0.015
0.041
0.080
0.171
0.289
1.710
391510024
monthly
all
130
4,131
0.019
0.046
0.000
0.001
0.001
0.002
0.003
0.008
0.018
0.040
0.071
0.131
0.196
1.123
290930021
monthly
1st quarter
127
1,057
0.018
0.042
0.000
0.001
0.001
0.002
0.003
0.007
0.017
0.040
0.061
0.137
0.211
0.583
290930021
monthly
2nd quarter
129
1,013
0.023
0.064
0.000
0.001
0.001
0.002
0.1X13
0.008
0.020
0.041
0.083
0.144
0.313
1.123
290930021
monthly
3rd quarter
129
1,040
0.017
0.031
0.000
0.000
0.001
0.002
0.003
0.008
0.017
0.040
0.064
0.105
0.139
0.520
391510024
monthly
4th quarter
128
1,021
0.019
0.041
0.000
0.001
0.001
0.002
0.003
0.008
0.017
0.039
0.075
0.135
0.182
0.633
290930016
3-month
all
129
4,116
0.019
0.037
0.000
0.001
0.001
0.002
0.004
0.009
0.019
0.043
0.071
0.126
0.209
0.534
290930021
3-month
1st quarter
127
1,053
0.019
0.035
0.000
0.001
0.001
0.002
0.003
0.008
0.018
0.043
0.066
0.143
0.208
0.361
290930016
3-month
2nd quarter
126
1,003
0.022
0.049
0.000
0.001
0.001
0.002
0.004
0.009
0.020
0.044
0.074
0.199
0.275
0.534
290930021
3-month
3rd quarter
128
1,037
0.019
0.034
0.000
0.001
0.002
0.002
0.004
0.009
0.018
0.041
0.074
0.108
0.152
0.396
391510024
3-month
4th quarter
126
1,015
0.018
0.029
0.001
0.001
0.002
0.002
0.004
0.009
0.019
0.044
0.073
0.105
0.137
0.324
290930016
N.sites = number of sites; N.obs = number of observations; SD = standard deviation; min = minimum; pi, p5, plO, p25,
p50, p90, p95, p98, p99 = 1st, 5th, 10th, 25th, 50th, 90th, 95th, 98th, 99th percentiles; max = maximum; max.site = AQS ID
number for the monitoring site corresponding to the observation in the max column. 1st quarter = January /February/March;
2nd quarter = April/May/June; 3rd quarter = July/August/September; 4th quarter = October/November/December.
Source = (U.S. EPA. 2023c).
Table 1-4 Distribution of Pb concentrations for various types of
measurements and monitoring site locations in pg/m3 for 2020-
2022
metric
measurement
network
N.sites
N.obs
mean
SD
min
pi
p5
plO
p25
p50
p75
p90
p95
p98
p99
max
max.site
daily
Pb-TSP
Source
83
16.901
0.030
0.095
0.000
0.000
0.001
0.002
0.003
0.008
0.021
0.061
0.114
0.238
0.363
2.923
290930021
daily
Pb-TSP
Non-Source
73
10.331
0.008
0.019
0.000
0.000
0.001
0.001
0-002
0.004
0.008
0.017
0.029
0.059
0.088
0.444
060250005
daily
Pb-PMlO
All Sites
48
7,625
0.003
0.008
0.000
0.000
0.000
0.000
0.001
0.002
0.003
0.005
0.008
0.016
0.025
0.319
420250300
daily
Pb-PM2.5
Urban
159
36.800
0.003
0.006
-0.012
-0.007
-0.005
-0.003
-0.001
0.002
0.005
0.009
0.012
0.016
0.019
0.360
060250005
daily
Pb-PM2.5
Rural
150
45,756
0.001
0.002
-0.009
0.000
0.000
0.000
0.000
0.000
0.001
0.001
0.002
0.004
0.007
0.099
511630003
monthly
Pb-TSP
Source
83
2,597
0.026
0.056
0.000
0.001
0.001
0.002
0.004
0.011
0.024
0.056
0.094
0.176
0.283
1.123
290930021
monthly
Pb-TSP
Non-Source
73
2,147
0.008
0.012
0.000
0.001
0.001
0.001
0.002
0.004
0.008
0.017
0.025
0.039
0.052
0.226
060250005
monthly
Pb-PMlU
All Sites
48
1,547
0.003
0.004
0.000
0.000
0.001
0.001
0.001
0.002
0.003
0.005
0.008
0.015
0.022
0.078
120250300
monthly
Pb-PM2.5
Urban
159
4,947
0.003
0.004
-0.008
-0.003
-0.001
0.000
0-001
0.002
0.004
0.006
0.008
0.011
0.014
0.077
060250005
monthly
Pb-PM2.5
Rural
151)
4,814
0.001
0.001
-0.001
0.000
0.000
0.000
0.000
0.000
0.001
0.001
0.002
0.004
0.006
0.015
511630003
3-month
Pb-TSP
Source
82
2,581
0.026
0.045
0.000
0.001
0.002
0.002
0.005
0.012
0.027
0.059
0.091
0.173
0.258
0.534
290930021
3-month
PlvTSP
Nou- Source
72
2,088
0.008
0.010
0.000
0.001
0.001
0.002
0.002
0.005
0.008
0.017
0.025
0.037
0.047
0.119
191550011
3-month
Pb-PMlO
All Sites
48
1.445
0.003
0.003
0.000
0.001
0.001
0.001
0.001
0.002
0.003
0.006
0.008
0.012
0.019
0.046
420250300
3-month
Pb-PM2.5
Urban
151
4,618
0.003
0.002
-0.002
-0.001
0.000
0.001
0.001
0.002
0.004
0.005
0.007
0.009
0.012
0.033
060250005
3-month
Pb-PM2.5
Rural
149
4,513
0.001
0.001
0.000
0.000
0.000
0.000
0.000
0.000
0.001
0.001
0.002
0.004
0.006
0.010
010730023
N.sites = number of sites; N.obs — number of observations; SD — standard deviation; rnin = minimum; pi, p5, plO, p25,
p50, p90, p95, p98, p99 = 1st, 5th, 10th, 25th, 50th, 90th, 95th, 98t.li, 99th percentiles; max = maximum; max.site = AQS
ID number for the monitoring site corresponding to the observation in the max column. Source — Source-Oriented Sites;
Non-Source = All Other Sites; Urban = CSN, NCore, and NATTS sites; Rural = IMPROVE sites.
Source = (U.S. EPA. 2023c).
Pb concentrations in ambient air in the United States have decreased since the 1970s, mainly due
to the phase-out of Pb in gasoline. In some cases, there has been a more recent period of continued
decline corresponding to reductions in Pb emissions from local and regional industrial sources. A
quantitative description of the trend based on monitoring network data is problematic for two reasons.
First, Pb concentration reporting requirements changed in 2010 from measured Pb concentration at
standard and temperature and pressure to Pb concentration measured concentration under local
conditions. As a result, daily concentration and design value data from before 2010 are not directly
1-64
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comparable to data from after 2010. Second, as numerous monitors have been discontinued because of
declining Pb concentrations, the proportion of monitors that are located near sources has increased.
Figure 1-9 is a national map and Figure 1-10 a time series plot showing how maximum rolling 3-month
average Pb concentrations in counties with Pb-TSP monitors changed during the periods 2010-2012 and
2020-2022 for all monitors that operated during both periods (U.S. EPA. 2023c). Pb concentrations
decreased for most monitors, in some cases by more than 0.05 |ig/nr\ Up-to-date graphics of annual
maximum 3-month average Pb concentration trends are available in "Overview of Lead (Pb) Air Quality
in the United States" (U.S. EPA. 2022a). and updated annually at https://www.epa.gov/air-qualitv-
analvsis/lead-naaqs-review-analvses-and-data-sets.
v Decreasing < 0.05 ug/mA3/yr (20 sites)
Source: (U.S. EPA. 2023c).
Figure 1-9 Site-level trends in maximum rolling 3-month average Pb
concentrations for 2010-2022.
1-65
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Edmonton (Bari and Kindzierski. 2016). and high traffic areas of Cincinnati (Grinshpun et al.. 2014). It is
also reflected in long-term sediment records from numerous locations in the United States, as described in
Section 1.2.3. In general, there was a steep decline in ambient air Pb concentrations in the 1970s and
1980s corresponding to the phase-out of Pb in gasoline, and in some cases a more recent period of
continued decline corresponding to reductions in Pb emissions from local and regional industrial sources.
In general, both AQS data and more detailed North American field studies usually support continuing
national and local long-term trends of decreasing ambient air Pb concentrations.
Source: (U.S. EPA. 2023c).
Note: Boxes represent the median and interquartile range, whiskers extend to the 1st and 99th percentiles, and values outside this
range are shown as circles. The red line shows the number of sites reporting regulatory data to the U.S. EPA in each year. The
concentrations on the left-hand y-axis are shown on a logarithmic scale.
Figure 1-11 Distribution of annual maximum 3-month concentrations
measured at regulatory Pb monitoring sites, 1980 to 2022.
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1.5.2.
Urban and Neighborhood Spatial Variability
The 2013 Pb ISA contains a comparison of Pb concentrations across six counties and past
literature on spatial distribution of airborne Pb in urban centers illustrating intra-urban variability (U.S.
EPA. 2013). These examples show that both point and non-point sources along with wind strength and
direction can play a role in distributing Pb across urban areas and create spatial variability. The 2006 Pb
AQCD (U.S. EPA. 2006) contains additional information on Pb transport in in the environment. Ambient
air Pb across urban and neighborhood scales may not be captured by national monitoring networks
because of contributions from more local emission sources (Yu et al.. 2011). In particular, high-emitting
Pb sources may create local hotspots of elevated Pb in environmental media (Section 1.3.4). Since the
publication of the 2013 Pb ISA, additional studies have been published illustrating Pb concentrations
across urban centers. Recent studies on near-road spatial variability are discussed together with traffic and
road sources in Section 1.2.6. More general aspects of urban and neighborhood spatial variability are
described in this section.
Spatial variability of ambient air Pb across New York City has been investigated in several
studies. A study of PM2.5 measurements across four boroughs of New York City found that spatial
patterns varied by season with the highest concentrations of Pb observed during the summer (7.94 ng/m3).
The sites within lower Manhattan had the highest concentrations Pb for both seasons. The authors
attributed Pb concentrations in lower Manhattan to waste incineration and traffic-related sources but
mention that Pb may be a poor tracer for incineration (Peltier and Lippmann. 2011). In a study of PM2.5
samples used in a land-use regression model from 150 street-level sites across New York City, Pb was
found to have a mean value of 3.40 ng/m3 with a standard deviation of 1.52 ng/m3 and spatial coefficient
of variation of 0.45. Pb was attributed most strongly to boilers burning residual oil, which was correlated
with other elements consistent with emission factors data, results from combustion experiments, and
characterization of residual oil fly ash (Ito et al.. 2016).
Other urban centers have also been investigated for spatial variability of ambient air Pb. Particles
between 2.5 and 10 |im were collected at ten sites in the greater Los Angeles area. The annual average Pb
concentration did not vary greatly between the Los Angeles (1.3 ng/m3), Long Beach (1.3 ng/m3),
Riverside (0.8 ng/m3), and Lancaster (0.5 ng/m3) sites. Principle component analysis revealed that Pb was
present most commonly with other elements that are tracers of abrasive vehicular emissions such as Sb,
Ba, Mo, Cu, Rh, and Fe (Pakbin et al.. 2011). In another study spatial variations of trace elements in PM10
around Paterson, NJ were investigated. Among 199 samples taken, there was an average Pb concentration
of 5.37 ng/m3 with a standard deviation of 5.07 ng/m3 (Yu et al.. 2011). In a study of trace metals
concentrations across four sites in St. Louis the authors found that annual median air Pb concentrations in
PM10 ranged from 6.01 ng/m3 (S.D. = 10.10 ng/m3) at a suburban site -10 km away from the urban core
to 8.96 ng/m3 (S.D. = 16.00 ng/m3) at an urban site 3 km north of the urban core. Conditional probability
function plot graphs revealed that all sites were affected by Pb from a source to the south. A local
smelting plant was identified as the possible source (Yadav and Turner. 2014).
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Stevens et al. (2014) investigated Pb-PlVb 5 concentrations across six sites in Detroit, Michigan
that were part of the Detroit Exposure and Aerosol Research Study. The authors found that Pb
concentrations in PM2 5 were heterogeneous across sites, with the highest mean concentrations of Pb for
outdoor, indoor, and personal exposures occurring at two heavily industrial areas focusing on steel
manufacturing and automobiles. In another study spatial variability of PM2 5 metals in Massachusetts was
modeled. Pb was estimated to be most heavily concentrated in areas of high roadway density in
downtown Boston, similar to areas high in Al, Fe and Ti, indicating that the source of Pb was likely from
road dust and soil particles. The mean value for Pb among 62 sites was 4.5 ng/m3 with a predicted
coefficient of variation of 0.467 (Rcquia et al.. 2019).
Upadhvav et al. (2011) investigated Pb in PM2 5 and PM10 around three sites in Phoenix, Arizona.
Two sites were located south of Phoenix Sky Harbor International Airport representing a mix of urban
residential and industrial use while a control site was located east of the airport with no local point
sources present. The two sites closest to the airport had the highest mean air Pb concentrations among the
three sites, 4.6 ng/m3 and 4.7 ng/m3 versus 2.0 ng/m3 for Pb-PMo.s and 6.3 ng/m3 and 5.6 ng/m3 versus
3.3 ng/m3 for Pb-PMio, similar to other U.S. cities. Principle component analysis revealed that Pb in PM2.5
was grouped with elements Cu and Zn, suggesting mobile sources and Pb in PM10 was grouped with As
and Cr, suggesting combustion processes as a source. Pb concentrations also peaked on January 1st,
suggesting the influence of local fireworks.
Researchers may also use biological organisms as bioindicators for Pb concentrations in ambient
air. Pigeons were used in a New York City study. Samples of pigeons" blood revealed that the highest Pb
concentrations were found in the Soho/Greenwich Village neighborhood (mean = 23.121 |ig Pb/dL) and
other neighborhoods of downtown Manhattan (Cai and Calisi. 2016). In another study, 26 lichen samples
and four atmospheric PM measurements were collected around Middletown, Ohio. Pb concentrations in
lichen samples reached a high of 151.27 ppm nearest the local steel plant but showed large heterogeneity
across samples with the lowest Pb concentration at 11.30 ppm found in a lichen sample outside the
general area, used as a background. Isotopic analysis of Pb species indicates that the Pb in these lichen
samples are a mix of coal fly ash and traffic-related PM, with some possible contribution from steel plant
emissions (Kousehlar and Widom. 2020).
Jovan et al. (2022) and Kondo et al. (2022) used local youth to collect moss samples around the
industrially adjacent South Park and Georgetown neighborhoods in Seattle, WA. Jovan et al. (2022)
found that the 79 samples collected by youth with minimal supervision had highly significant agreement
(p = 0.001) to 19 expert-collected samples. Pb concentrations peaked along the industrial core separating
the two communities. Among all samples there was a median value of 18.1 mg Pb/kg and minimum and
maximum values of 5.9 and 110.6 mg Pb/kg respectively. Kondo et al. (2022) assessed the spatial
predictors of metal concentrations found in 61 of the moss samples. Traffic volume and block-group level
percent people of color were found to be the spatial predictors significantly associated with higher Pb
concentrations in moss.
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1.5.3. Seasonal and Diurnal Trends
The 2013 Pb ISA (U.S. EPA. 2013) briefly illustrates that, depending on the measurement
location, there can be seasonal variation for Pb in ambient air. Seasonal variability can depend heavily on
local meteorological conditions including mixing height, wind direction, precipitation, and humidity (U.S.
EPA. 2006). Levin et al. (2020) discusses several additional factors that may contribute to local trends in
ambient air Pb. As explored in Section 1.3.4, resuspension of Pb in soil can contribute Pb into air, with
the highest contributions in dry summer months. Wildfires, which occur most often during the summer
and into fall, can remobilize Pb deposited in the natural environment and Pb contained in man-made
structures (Section 1.2.4). The 2013 Pb ISA (U.S. EPA. 2013) includes past research that has identified
seasonal variation of ambient air Pb in various locations. There was no research captured by the literature
screening for this document that contained information on ambient air Pb diurnal trends.
Seasonal variation of Pb concentrations has been investigated in several studies of various
locations since the 2013 Pb ISA (U.S. EPA. 2013). These studies have varied in their design with some
measuring trends over a period of several years while others only measure a one-year period, and some
have presented averages of Pb concentrations while others have presented more detailed monthly data of
Pb concentrations. Table 1-5 below details study conditions and findings of seasonal variations from these
studies.
Table 1-5
Seasonal variations in Pb concentration in ambient air
Study
Location
Time
Period
Seasons
(Months)
Source
Attribution of
Pb
Findings of Seasonal Variations
(Pakbin et
al., 2011)
Greater
Los
Angeles
Area
2008-
2009
All Seasons
Abrasive
vehicular
emissions with
some
contributions
from soil dust
and vehicular
catalytic
converter wear
Average ambient air Pb
concentrations highest in the
winter for the more urban Los
Angeles (1.3 ng/m3 in winter
versus 1.0 ng/m3 in summer) and
Long Beach (1.8 ng/m3 in winter
versus 0.8 ng/m3 in summer) sites.
Average ambient air Pb
concentrations were higher in the
summer for the semirural Riverside
(0.5 ng/m3 in winter versus
0.8 ng/m3 in summer) and desert
Lancaster (0.2 ng/m3 in winter
versus 0.7 ng/m3 in summer) sites.
(Peltier and
Lippmann,
2011)
New York
City, NY
Winter (January-
2008 March), Summer
(May-July)
Incineration and
Biomass Burning
Highest concentrations of Pb
measured were found in lower
Manhattan during the summer
months, suggesting a highly
localized source of Pb only present
during this time. Pb concentrations
were near 1 ng/m3 with hot spot
concentrations, on average, at 5
and 10 ng/m3 during the winter and
summer, respectively.
1-70
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Study
Location
Time
Period
Seasons
(Months)
Source
Attribution of
Pb
Findings of Seasonal Variations
(Sona and
Gao. 2011)
Carlstadt,
NJ
2007-
2008
Winter
(December-
February),
Summer (July)
Brake wear,
direct vehicle
emissions, other
urban sources
Higher concentrations of Pb in
winter for both coarse (2.04 ng/m3
in winter versus 1.41 ng/m3 in
summer) and fine (2.82 ng/m3 in
winter vs. 1.27 ng/m3 in summer)
despite overall aerosol mass being
lower in winter than summer.
(Yu et al..
2011)
Paterson,
NJ
2005-
2006
All Seasons
Traffic-related
and industrial
emissions
Pb concentrations highest in the
winter and second highest in the
fall for mobile, industrial, and
commercial sites. Pb
concentrations were highest in
winter and summer for the
background site.
(Grinshpun
et al., 2014)
Cincinnati,
OH
2010-
2011
Fall, Winter,
Summer
(months not
specified)
Traffic-related
emissions
Pb concentration was higher in the
fall than winter and summer in a
downtown Cincinnati location.
(Kundu and
Stone.
2014)
Iowa
2009-
2012
All Seasons
Diesel
combustion,
gasoline
combustion,
biomass burning,
industry
There were no consistent seasonal
patterns for Pb concentrations
across sites measured.
(Prabhakar
et al., 2014)
Southern
Arizona
1988-
2009
Summer (April-
June), Fall
(October-
November),
Monsoon (July-
September),
Winter
(December-
March)
Smelting
operations and
traffic-related
emissions
Seasonal patterns were found to
vary across sites, indicative of
local conditions. Phoenix was
impacted more heavily by traffic-
related emissions, being an urban
center, while Tonto National
Monument was more affected by
related smelting operations.
(Stevens et
al.. 2014)
Detroit,
Michigan
2004-
2007
Summer, Winter
(months not
specified)
Traffic-related
and industrial
emissions
Outdoor mean Pb concentrations
were found to be slightly higher in
summer for four sites. In contrast,
site 5, a heavily industrial site had
much a higher average Pb winter
value (42 ng/m3) than summer
value (15 ng/m3).
1-71
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Study
Location
Time
Period
Seasons
(Months)
Source
Attribution of
Pb
Findings of Seasonal Variations
(Li and
McDonald;
Gillespie.
2020)
Tulsa, OK
and
Picher, OK
2010-
2016
All Seasons
Atmospheric Pb
blown from
nearby chat piles
Picher showed strong seasonal Pb
mass concentration peaks during
the period of January-March and
September, likely due to Pb blown
from nearby chat piles. Tulsa did
not show any strong seasonal
variation over the measurement
period.
NJ = New Jersey; NY = New York; OH = Ohio; OK = Oklahoma; Pb = lead.
1.5.4. Particle Size Characteristics
The size distribution of Pb-containing particles differs depending on source type and the
collection efficiency of Pb-TSP samplers. Cho et al. (2011) found that most studies included in their
review and published after 1986 indicated a shift in Pb particle size distribution from the fine fraction to
coarse fraction with the primary mode rising to 2.5-10 |im from a previous estimation of it being below
2.5 |im. Studies used to evaluate this shift ranged from sampling near roads, near industrial sources,
offshore in a lake environment, rural locations, and urban locations, within the United States and the
European Union. The elimination of the combustion of tetramethyl- and tetraethyl Pb in automobiles as
the dominant source of Pb-PM in the atmosphere, as indicated in Section 1.2, has led to larger Pb-
containing particles on average. However, coarse particles have higher settling velocities than fine
fraction or ultrafine fraction particles, resulting in measured concentrations of coarse or ultracoarse
(particles greater than 10 (mi diameter) particles being spatially and temporally heterogeneous, as these
particles may drop out before they are collected by a TSP sampler. These topics have been discussed in
previous assessments (U.S. EPA. 2013. 2006) as well as in Sections 1.3.1.1 and 1.4.3.
The literature search and screening process for the current iteration of the ISA did not find many
published studies containing information on Pb particle size distributions beyond what was included in
the 2013 Pb ISA (U.S. EPA. 2013). To briefly summarize information from the 2013 Pb ISA, near-
roadway emissions may be the result of emissions directly from vehicular combustion or parts such as
brakes or wheel weights, sources near the roadway that are not related to traffic, or traffic-induced
turbulence causing resuspension of deposited Pb-bearing particles originating from wheel weights,
industrial emissions, or historic sources into air. Sabin et al. (2006) mentioned that freeways tend to be a
source of very large particles dispersed by turbulence from vehicular traffic and found a bimodal
distribution of particles at a near-road site in Los Angeles with a mode in the largest size fraction
sampled.
Other near-road studies in the 2013 Pb ISA (U.S. EPA. 2013) found a mix of size distributions at
near-road sites subject to different meteorological conditions and measurement techniques. Havs et al.
(2011) measured Pb in ambient air particles 20 meters north of a major interstate in Raleigh, NC. Pb was
1-72
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found in PM10-2.5, PM2.5-0.1 and PM0.1 size fractions, at 50 mg Pb/kg in each size fraction. Ambient air Pb
concentrations appeared unimodal and normally distributed over the accumulation mode with
0.4 ± 0.4 ng/m3, 1.4 ± 0.6 ng/m3, and 0.1 ± 0.02 ng/m3 for PM10-2.5, PM25 u i. and PM0.1 size fractions,
respectively. Daily concentration changes were heavily correlated with traffic, including Pb-PMio samples
highly correlated with As samples, most likely resuspended from contemporary roadway sources. Song
and Gao (2011) collected measurements of ambient air PM using a sampling site approximately 5 meters
from the New Jersey Turnpike. The Pb mass size distribution had a bimodal concentration distribution in
summer and a trimodal distribution in winter with 47% and 58% of Pb mass measured in fine particles in
summer and winter, respectively (Pb mass concentration values mentioned in Table 1-5). Factor analysis
attributed the source of Pb to brake wear, fuel combustion, and urban pollution.
Masri et al. (2015) collected both fine-mode and coarse mode ambient air PM at the Harvard
supersite in Boston, MA. This site is located atop a six-story building within one block of a four-lane
roadway and two major highways. Trace amounts of Pb were found to be exclusively associated with
fine-mode particles. Positive Matrix Factorization analysis was used to associate these particles across a
wide range of possible sources from regional pollution, motor vehicles, crustal or road dust, oil
combustion and wood burning.
Size distributions have also been recorded at other sites as well, reflecting spatiotemporal
variability within and near cities. Gonzalez et al. (2021) analyzed data from five sites (Manila,
Philippines; Marina, CA; Tucson, AZ; Hayden, AZ; Mt. Lemmon, AZ) that measured both
submicrometer and supermicrometer particles (range 0.056-18.0 (mi) which were then extracted for
further analysis. Pb mass concentrations within the submicrometer and supermicrometer ranges were
found to vary by site with Manila, Tucson, and Hayden (the location of an active metals smelter) having
higher mass concentrations in the submicrometer range while Marina and Mt. Lemmon had higher Pb
mass concentrations in the supermicrometer range. The Marina and Manila sites were also separated by
fire and non-fire-influenced datasets which showed the presence of a submicrometer mode for Pb mass
concentration in the fire-influenced data that was not present in the non-fire-influenced data. Upadhvav et
al. (2011) measured ambient air PM in Phoenix, Arizona both south and east of the Phoenix Sky Harbor
International Airport. The authors found Pb to be associated with both fine and coarse particles at three
sites in Arizona, with PM2.5:PMio ratios between 0.5 and 0.7 for Pb at the three sites tested, indicating that
a substantial fraction of Pb was associated with both fine and coarse particles at each site. Youn et al.
(2016) performed chemical speciation on aerosol samples collected at an active smelting site in Hayden
AZ and an urban background site in Tucson AZ. The particle size distribution of Pb mass found at the
active smelting site was bimodal with a large peak at 0.32 |im. and a smaller peak at 5.6 (mi. The
background site was trimodal with peaks at 0.1 |im. 0.32 (mi, and a smaller peak at 3.2 (mi.
Submicrometer particles were attributed to the condensation and coagulation of smelting vapors, whereas
coarse particles were attributed to fugitive dust, including from mine tailings. Ambient air Pb
concentration values at the active smelting site were higher than in samples from the background site.
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1.5.5. Background Concentrations
A small fraction of Pb in ambient air in the United States cannot be reduced by domestic emission
controls or domestic interventions within the United States. According to the 2013 Pb ISA, natural
sources of Pb to ambient air include suspension of surface soil containing natural Pb and wind erosion of
natural Pb-containing rocks. Previous assessments have suggested the evidence to indicate a plausible
range of natural background Pb concentration is 0.02 to 1 ng/m3 (U.S. EPA. 2013; NRC. 1980). As
described in the 2013 Pb ISA (U.S. EPA. 2013). average concentrations in this range have been measured
at remote monitoring sites such as Crater Lake, OR and Lassen Volcanic National Park, CA. In addition
to natural sources, intercontinental transport of Asian dust could also make a substantial contribution to
total atmospheric Pb, but generally less than 1 ng/m3 (U.S. EPA. 2013). The 2013 Pb ISA reviewed
evidence for intercontinental transport of Pb in African dust to the Southeastern United States and
described mixed results concerning whether intercontinental transport or natural sources contributed the
most to atmospheric Pb. The 2013 Pb ISA concluded that estimates of background Pb concentrations in
ambient air were well below current concentrations (U.S. EPA. 2013). No new studies on background Pb
concentrations in the United States since the last NAAQS review were identified in our literature search
and screening process.
1.6 Summary and Conclusions
Pb emissions and ambient air concentrations in the United States continue to steadily decline.
Major industrial sources have either reduced their emissions or closed, resulting in the emergence of
aviation gas as the dominant contemporary source. However, resuspension of soil containing Pb from
legacy sources has the potential to contribute to atmospheric Pb in some locations, and high
concentrations of Pb associated with wildfires have been observed. Pb from these sources continues to
have potential health and ecological effects after atmospheric deposition to soil and water. A substantial
fraction of airborne Pb can be associated with PM larger than 10 |im in some locations. Bias and
uncertainty associated with sampling these large particles in Pb-TSP sampling are still an issue, although
it has become better understood and there have been several improvements in measurement tools,
including development of an ICP-MS-based FRM for Pb analysis, and introduction of reference materials
for analysis on filters. Overall, there have been substantial improvements in our understanding of and
research capabilities for airborne Pb.
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1.7 References
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review [Review]. Sci Total Environ 599-600: 1740-1755. http://dx.doi.Org/10.1016/i.scitotenv.2017.05.096.
Alexakis. DE. (2020). Suburban areas in flames: Dispersion of potentially toxic elements from burned vegetation
and buildings. Estimation of the associated ecological and human health risk. Environ Res 183: 109153.
http://dx.doi.Org/10.1016/i.envres.2020.109153.
Alkhatib. E. id: Akhdhar. A: Altirad. F. (2015). Prediction of metal remobilization from sediments under various
physical/chemical conditions: Design of experiment for Cd, Co, Pb, and Zn [Abstract]. Abstr Pap Am
ChemSoc 249: 141.
Alshettv. D: Shiva Nagendra. SM. (2022). Urban characteristics and its influence on resuspension of road dust, air
quality and exposure. Air Qual Atmos Health 15: 273-287. http://dx.doi.org/10.1007/sl 1869-021-01102-x.
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United States
Environmental Protection
Agency
EPA/600/R-23/375
January 2024
www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 2: Exposure, Toxicokinetics, and
Biomarkers
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
-------
CONTENTS
DOCUMENT GUIDE 2-iii
LIST OF TABLES 2-v
LIST OF FIGURES 2-vi
ACRONYMS AND ABBREVIATIONS 2-vii
APPENDIX 2 EXPOSURE, TOXICOKINETICS, AND BIOMARKERS 2-1
2.1 Exposure 2-1
2.1.1 Overview of Pathways for Pb Exposure 2-2
2.1.2 Environmental Exposure Assessment Methodologies 2-4
2.1.3 Exposure Studies 2-6
2.1.4 Co-Contaminants Commonly Present with Pb 2-30
2.1.5 Exposure Disparities for Specific Populations 2-33
2.2 Kinetics 2-42
2.2.1 Absorption 2-43
2.2.2 Distribution and Metabolism 2-56
2.2.3 Elimination 2-65
2.3 Pb Biomarkers 2-67
2.3.1 Bone-Pb Measurements 2-67
2.3.2 Blood-Pb Measurements 2-68
2.3.3 Urine-Pb Measurements 2-69
2.3.4 Pb in Other Biomarkers 2-69
2.3.5 Relationship between Pb in Blood and Pb in Bone 2-72
2.3.6 Relationship between Pb in Blood and Pb in Soft Tissues 2-83
2.4 Studies of Pb Biomarker Levels 2-87
2.4.1 Pb in Blood 2-87
2.4.2 Pb in Bone 2-98
2.4.3 Pb in Urine 2-99
2.4.4 Pb in Other Biomarkers 2-101
2.5 Empirical Models of Pb Exposure-Blood Pb Relationships 2-101
2.5.1 Air Pb-Blood Pb Relationships in Children 2-103
2.5.2 Air Pb-Blood Pb Relationships in Adults 2-111
2.5.3 Soil Pb-Blood Pb Relationships 2-113
2.6 Biokinetic Models of Pb Exposure-Blood Pb Relationships 2-116
2.7 Summary and Conclusions 2-119
2.7.1 Exposure 2-119
2.7.2 Toxicokinetics 2-120
2.7.3 Pb Biomarkers 2-122
2.7.4 Air Pb-Blood Pb Relationships 2-123
2.8 References 2-124
2-iv
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LIST OF TABLES
Table 2-1 Comparison of personal, indoor, and outdoor Pb-PM measurements from several studies
included in the 2013 Pb ISA 2-7
Table 2-2 Pb-PIVh 5 concentrations across six sites in Detroit, Michigan 2-8
Table 2-3 Median soil Pb concentrations in New Orleans census tract level surveys 2-12
Table 2-4 Dietary exposures to Pb based on U.S. Food and Drug Administration Total Diet Study
(2014-2016) and What We Eat in America (2009-2014) food consumption data 2-19
Table 2-5 Contribution of maternal blood Pb to breast milk at 1-3 months postpartum 2-24
Table 2-6 Pb content in various consumer products 2-25
Table 2-7 Co-contaminants in Pb sources 2-32
Table 2-8 Specific unique combinations of As, Cd, Pb, and Hg detected at or above the respective
median concentrations in urine or blood among the U.S. population 6 years and older,
National Health and Nutrition Examination Survey 2007-2012 data 2-33
Table 2-9 Prevalence of elevated blood Pb levels in refugee children 2-37
Table 2-10 Relative bioavailability for varied Pb forms and sources 2-52
Table 2-11 Blood-Pb concentrations in the U.S. population 2-87
Table 2-12 Urine-Pb concentrations in the U.S. population 2-99
Table 2-13 Summary of estimated slopes for blood Pb-to-air Pb slope factors in children 2-105
2-v
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LIST OF FIGURES
Figure 2-1 Conceptual model of air-related Pb exposure through inhalation and ingestion. 2-3
Figure 2-2 Distribution of Pb in road dust samples collected in three industrial and mining towns
located in southern Poland. 2-56
Figure 2-3 Plot of blood and plasma Pb concentrations measured in adults and children. 2-58
Figure 2-4 Relationship between Pb intake and blood Pb concentration in infants (n = 105, age
13 weeks, formula fed). 2-59
Figure 2-5 Simulation of quasi-steady state blood and plasma Pb concentrations in a child (age
4 years) associated with varying Pb ingestion rates. 2-61
Figure 2-6 Simulation of relationship between blood Pb concentration and body burden in children,
with an elevated constant Pb intake from age 2 to 5 years. 2-76
Figure 2-7 Half-times of Pb in blood as reported by Specht et al. (2019b). 2-77
Figure 2-8 Simulation of relationship between blood Pb concentration, bone Pb, and body burden in
adults. 2-80
Figure 2-9 Simulation of blood and soft tissue (including brain) Pb in children and adults who
experience a period of increased Pb intake. 2-84
Figure 2-10 Simulation of blood and brain Pb in children and adults who experience a period of
increased Pb intake. 2-85
Figure 2-11 Relationship between Pb in urine, plasma, blood, and bone. 2-86
Figure 2-12 Temporal trend in blood Pb concentrations. 2-89
Figure 2-13 Blood Pb cohort means versus year of exam. 2-90
Figure 2-14 Blood Pb geometric means versus year of NHANES exam by race/ethnicity. 2-90
Figure 2-15 Geometric mean childhood blood Pb levels assessed between 1 and 8 years old, stratified
by race/ethnicity. 2-93
Figure 2-16 Slope factors for blood Pb as a function of air Pb. 2-108
Figure 2-17 Blood Pb versus soil Pb for two New Orleans surveys completed in 2001 and 2017. 2-114
Figure 2-18 Comparison of slope factors in New Orleans data on linear-linear (top) and log-log
(bottom) plots. 2-115
2-vi
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ACRONYMS AND ABBREVIATIONS
(x-SRXRF microbeam synchrotron radiation X-ray
fluorescence
1OOLL 100 octane, low lead
AALM All-Ages Lead Model
AAS atomic absorption spectrometry
AERMOD American Meteorological
Society/Environmental Protection Agency
Regulatory Model
Ag silver
AF absorbed fraction
AHHS American Healthy Homes Survey
A1 aluminum
ALAD S-aminolevulinic acid dehydratase
ALF artificial lysosomal fluid
ALM Adult Lead Methodology
As arsenic
AQCD Air Quality Criteria Document
AQS Air Quality System
ASV anodic stripping voltammetry
Ba barium
BLL blood lead level
BLRV blood lead reference value
BMI body mass index
Ca calcium
Cal EPA California Environmental Protection Agency
Cd cadmium
CDC Centers for Disease Control and Prevention
Co cobalt
Cr chromium
CR creatine
CSMR chloride to sulfate mass ratio
Cu copper
DoD Department of Defense
DOE Department of Energy
DOHMH Department of Health and Mental Hygiene
(New York City)
DWSD Detroit Water and Sewage Department
S-ALA S-aminolevulinic acid
EBLL elevated blood lead level
FDA Food and Drug Administration
FWSC Flint Water Service Center
Ga gallium
Ge germanium
GFR glomerular filtration rate
GI gastrointestinal
GM geometric mean
GSD geometric standard deviation
HA Housing Authority
HC hydrocarbon
HFE hemochromatosis gene
Hg mercury
ICP-AES inductively coupled plasma atomic emission
spectroscopy
ICP-MS inductively coupled plasma mass
spectrometry
ICRP International Commission on Radiological
Protection
IDF Israeli Defense Forces
IEUBK Integrated Exposure Uptake Biokinetic
IMPROVE Interagency Monitoring of Protected Visual
Environments
In indium
IRL interim reference level
ISA Integrated Science Assessment
IVBA in vitro bioaccessibility
K-XRF K-shell X-ray fluorescence
La lanthanum
LA-ICP-MS laser ablation-inductively coupled plasma-
mass spectrometry
LOD limit of detection
LSL lead service line
LTC loading to concentration
Mg magnesium
MG Mahayogaraj Guggulu
MMB multimedia biomarker
mo month(s)
Mn manganese
NAAQS National Ambient Air Quality Standards
NATA National Air Toxics Assessment
Ni nickel
NR not reported
N.D. not detected
NHANES National Health and Nutrition Examination
Survey
NHEXAS National Human Exposure Assessment
Survey
OSHA Occupational Safety and Health
Administration
PAH polycyclic aromatic hydrocarbon
Pb lead
PbA air Pb concentration
PbB blood Pb concentration
PIR poverty-income ratio
2-vii
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PM particulate matter TDS
PUFA polyunsaturated fatty acids Ti
RBA relative bioavailability TRI
RBC red blood cell TSP
RSD relative standard deviation U.S. EPA
S sulfur
SAB Scientific Advisory Board ^
SD standard deviation ^A
Se selenium WQS
SES socioeconomic status WWEIA
SHEDS Stochastic Human Exposure and Dose XRF
Simulation yr
SLL soil lead level Zn
Sm samarium Zr
Sn tin
Sr strontium
Total Diet Study
titanium
Toxics Release Inventory
total suspended particles
United States Environmental Protection
Agency
vanadium
Veterans Affairs
weighted quantile sum
What We Eat in America
X-ray fluorescence
year(s)
zinc
zirconium
2-viii
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APPENDIX 2 EXPOSURE, TOXICOKINETICS,
AND BIOMARKERS
The purpose of this appendix is to review exposure, toxicokinetic, and biomarker information
relevant to human lead (Pb) exposure, with a focus on scientific literature from 2011 onward. Section 2.1
reviews pathways of Pb exposure, exposure assessment methodologies, exposure studies by various
pathways, co-exposures, and exposure disparities for specific populations. Section 2.2 reviews absorption,
distribution and metabolism, and elimination of Pb from the body. Section 2.3 reviews methodologies for
biomarker measurement and the relationships between blood Pb and Pb in bone and soft tissues.
Section 2.4 reviews studies of biomarker levels, including trends in Pb biomarker levels over time.
Sections 2.5 and 2.6 review empirical models and biokinetic models of Pb exposure - Pb blood
relationships, respectively. Section 2.7 presents overall conclusions on the scientific evidence reviewed
within this appendix.
2.1 Exposure
The purpose of this section is to review studies, with a focus on recent literature, that provide
information about human exposure to Pb through the environment. Because Pb body burden is often used
to estimate exposures (e.g., Pb concentrations in blood, bone, etc.), and because air-related Pb exposure
may occur through inhalation or ingestion of materials that have been contaminated by Pb originally
found in ambient air, this appendix evaluates the evidence for total Pb exposures, including inhalation
exposures and exposures from ingestion of food, water, dust and soil, and other materials. Lack of data
makes it a challenge to trace Pb to air in biomarker studies using speciation or isotopic signatures.
The information in this chapter builds on conclusions from the 2013 Pb Integrated Science
Assessment (hereinafter referred to as the 2013 Pb ISA) (U.S. EPA, 2013), which found that air Pb
concentrations and blood Pb levels (BLLs) have continued to decrease over the past 45 years. The
phasing out of leaded gasoline and reductions in point source Pb emissions have been important
contributors to this decline. Section 1.5.2 of this current ISA reports the national median of the annual
maximum 3-month average Pb concentration declined by 88% from 2010 to 2021. Section 1.5.2 contains
more details on national Pb air concentration temporal trends (https://assessments.epa.gov/isa/document/
&deid=359536).
As described in detail in Section 2.4.1, there has been a decline in BLLs from 1976 to 2018 in all
birth cohorts. The geometric mean (GM) BLL across all subjects surveyed in the 1999-2000 National
Health and Nutrition Examination Survey (NHANES) cycle was 1.66 (ig/dL (95% CI: 1.60, 1.72). The
GM BLL across all subjects surveyed in the 2017-2018 NHANES cycle was 0.753 (ig/dL (95% CI:
0.723, 0.784). BLLs have decreased among all age and race/ethnicity groups. GM BLL differences
between non-Hispanic Black children and other racial/ethnic groups have also lessened over time.
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Despite the drop in air Pb concentrations and human BLLs overtime, sources of Pb still remain.
This section discusses exposure to Pb in air and other environmental media, including soil, dust, and
water. It also discusses Pb exposure through other pathways, including diet, consumer products,
ammunition, and occupational exposures. Co-exposures and exposures in specific populations are also
briefly discussed.
2.1.1 Overview of Pathways for Pb Exposure
Since the publication of the 2013 Pb ISA (U.S. EPA. 2013). the environmental pathways for Pb
exposure have remained consistent, whereas the amounts of Pb from various sources have changed. Pb
has multiple point and nonpoint sources and passes through various environmental media, including air
(the focus of this assessment), soil, or water.
The diagram (Figure 2-1) below depicts the various pathways that ambient air Pb can take
through the environment to reach a human being. Exposures are considered air-related if they pass
through the air compartment at any point prior to plant, animal, or human contact. For example, air-
related Pb exposure may occur through inhalation or ingestion of food, water, dust and soil, or other
materials that have been contaminated by Pb originally in ambient air. Additionally, organisms can be
exposed to Pb directly through contact with air that contains Pb. Non-ambient air-related exposures
include those from an occupation, hand-to-mouth contact with Pb-containing consumer goods, hand-to-
mouth contact with dust or chips of peeling Pb-containing paint, or ingestion of Pb in drinking water
conveyed through Pb pipes. Pb body burden is an aggregation of all of these different exposures.
Pb in the ambient air is found in particles that vary in size depending on the emission source and
whether there is entrainment of environmental media, such as resuspension of soil. Pb-containing
particulate matter (Pb-PM) emitted from automobiles and piston engine aircraft have been found to be
smaller than those emitted by industrial sources, resuspended soil, and tire/break wear (Griffith. 2020;
U.S. EPA. 2013; Schauer et al.. 2006). Lee et al. (1972) found, in annual data collected before 1976, that
Pb-PM in several urban areas dominated by traffic sources consisted of primarily small particles (59%-
74% <1 (mi; 74%-87% <2 (mi). Locations near industrial sources or impacted by resuspended road dust
typically show the lowest Pb PM2.5/PM10 ratios and the highest Pb-TSP or Pb-PMio concentrations (Cho
et al.. 2011).
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This figure displays air-related exposure pathways of Pb through the environment. Dashed lines represent resuspension of Pb into
the air. Green ovals represent sources of Pb not associated with the air compartment but may contribute to Pb along the exposure
pathway. Pb from processing includes Pb that may end up in diet as a result of intentional or inadvertent addition of Pb to food or
food additives such as spices. Other recognized sources of Pb exposure such as occupational or some consumer products are not
ambient air-related and as such are not included in this figure.
Figure 2-1 Conceptual model of air-related Pb exposure through inhalation
and ingestion.
Concentrations of Pb in an individual's microenvironment and that individual's time spent doing
different activities can influence their exposure. The importance of different sources and pathways of Pb
exposure varies across the U.S. population and is situation specific. As an example, Pb in soil was found
to likely be a key pathway of exposure for children in pre- and post-Katrina New Orleans, whereas Pb-
contaminated drinking water was a key contributor to elevated BLLs (EBLLs) during the Flint Water
Crisis (Gomez et al.. 2018; Mielke et al.. 2017; Zahran et al.. 2017b; Kennedy et al.. 2016). Evidence
points to incidental ingestion of Pb dust by hand-to-mouth activity as a leading exposure route for young
children (Mielke et al.. 2017; von Lindern et al.. 2016; U.S. EPA. 2013).
Frank et al. (2019) performed a meta-analysis looking at Pb measured in multiple environmental
media over the last 20 years. The mean estimate of Pb in urban residential soils was three times higher
than in rural residential soils. The infrastructure in urban centers tends to be older than rural residential
communities, and those urban centers are closer to Pb sources including dense traffic networks, industrial
emissions, and brownfield sites. Data were limited for most environmental media, with soil and dust
being the most robust in terms of the amount of available literature. More comparisons across
environmental media with consistent sampling methodology would be worthwhile to understand the
contribution of each environmental pathway to exposure.
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2.1.2
Environmental Exposure Assessment Methodologies
Various monitoring techniques are used to estimate exposure to Pb from the environment. The
2013 Pb ISA (U.S. EPA, 2013) contains brief descriptions of some of these techniques, and the 2006 Pb
Air Quality Criteria Document (AQCD) (U.S. EPA, 2006) contains more detailed information. To
understand the contributions of particular pathways to overall Pb exposures, measurements in air, soil,
and dust are performed. Ambient air monitoring techniques are described in detail in Section 1.4
(https://assessments.epa.gov/isa/document/&deid=359536). Four national monitoring networks (State or
Local Air Monitoring Stations, Chemical Speciation Network, Interagency Monitoring of Protected
Visual Environments [IMPROVE], and National Air Toxics Trends Station) collect data on Pb
concentrations and report data to the United States Environmental Protection Agency's (U.S. EPA's) Air
Quality System (AQS) (U.S. EPA, 2019a). National Ambient Air Quality Standards (NAAQS)
compliance must be determined using Federal Reference Methods or Federal Equivalent Methods that
measure Pb in total suspended particles (TSP). In some studies, indoor and personal monitoring of Pb
have also been performed to understand these Pb concentrations" relationships to outdoor air and possible
exposure (e.g., Stevens et al. (2014)).
Pb in soil can be measured using inductively coupled plasma atomic emission spectroscopy (ICP-
AES) of a nitric acid digested sample (Wharton et al„ 2012), inductively coupled plasma mass
spectrometry (ICP-MS) (Yu et al„ 2022), or atomic absorption spectrometry (AAS) (Okonkwo et al.,
2021). It has also been measured by U.S. EPA method 6200, which uses portable X-ray fluorescence
(XRF) (Obcng-Gvasi et al., 2021). Dust samples for Pb analysis are collected using wipe sampling and
vacuum sampling as described in the 2013 Pb ISA (U.S. EPA, 2013), and these techniques have not
drastically changed since they were first formalized.
The multiple pathways by which individuals are potentially exposed to Pb can make it
challenging to determine the Pb sources causing an individual's BLL to become elevated. There is often
no single primary source on the individual level as all exposures contribute to Pb body burden. The ratio
of three isotopes (2ll6Pb, 2ll7Pb, and 2ll8Pb) in human biomarkers can be compared with those found in
indoor, outdoor, and occupational sources to provide supporting evidence of where Pb originated (Jaeger
et al., 1998). Isotopic analysis is a research tool and most often used in studies designed to investigate Pb
sources for a certain population or identify sources of Pb that contaminate soil/dust. Becker et al. (2022)
used isotope ratios (2u8Pb/2"6Pb and 2u6Pb/2"7Pb) to apportion the sources of blood Pb in five children (ages
1-6 years) living in urban areas of Kansas City, MO, who were screened for EBLLs (>5 (ig/dL). Indoor
dust samples were collected in play areas where the children spent significant time. One child's blood Pb
was isotopically similar to Pb in both indoor dust and yard soil, which likely contributed to the indoor
dust Pb based on the similarity in their isotopic ratios. Turmeric, indoor dust, and paint chips were each
individually identified as a dominant source of Pb in blood for three other children. For a child having the
highest BLL, nearly 13 (ig/dL, the isotopic Pb ratios in blood were dissimilar to the Pb ratios in yard soil,
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indoor dust, and paint chips. This suggests that another unsampled Pb source in the home or elsewhere
was causing this child's EBLL.
Xue et al. (2022) developed a generalizable approach to identify locations of hotspots for Pb
exposure based on children's elevated BLLs and analyzed Pb models or indices as surrogates of exposure
in those locations. A case study was used to apply the approach. BLL data (1,930,943 samples) for
children <6 years of age were obtained from the Michigan Department of Health and Human Services for
the 2006-2016 period. The BLL data (half venous, half capillary samples) was well correlated (r=0.58)
between venous and capillary samples within person and year. These BLL data were geocoded to
Michigan census tracts. A top 20-percentile method and geospatial cluster analysis method were both
used to identify census tracts with high %EBLLs. In addition, U.S. EPA's EJSCREEN 2017 Pb Paint EJ
Index, modeled BLLs of Schultz et al. (2017). and the U.S. Department of Housing and Urban
Development's Deteriorated Paint Index were used as Pb models/indices for analyzing old housing data
and sociodemographic variables. These were analyzed and mapped at census tract resolution with model-
to-model comparison by analyzing the models against one another and modeling them against %EBLL
data to see if they are useful as surrogates in absence of BLL data. The percentage of census tracts in
Michigan with an exceedance rate > 10% for BLLs > 5 (ig/dL decreased from 14.8% in 2006-2007 to
4.1% in 2014-2016. The three Pb models/indices had high statistical convergence, indicating high
similarity between them. Both the geospatial clustering approach and the 20-percentile method had
moderate statistical convergence with %EBLLs from 2014 to 2016. The method was found to be able to
inform hotspot identification for Michigan, however, the authors acknowledged that this method may
miss some locations as the methods did not identify all areas of EBLLs in this study and should be
verified by available blood Pb data and information about local Pb sources and exposures. Zartarian et al.
(2022) provides a broader state-of-the-science overview of Pb geospatial mapping approaches including
links to publicly available BLL data from 32 state health departments, a multimedia environmental
sources data table, and a summary of available Pb exposure indices and their data.
In addition to using measurements of Pb concentrations in environmental media and biomarkers,
various modeling strategies have been used to estimate Pb exposure. Air dispersion models estimate the
spread of Pb releases throughout the air in a certain region. For example, the American Meteorological
Society/Environmental Protection Agency Regulatory Model (AERMOD) is a steady-state plume model
that considers short-range dispersion from stationary industrial sources in the planetary boundary layer
over both simple and complex terrain (Cimorelli et al.. 2005; Perry et al.. 2005). Several studies have used
AERMOD to estimate air Pb concentrations around industrial facilities (e.g., Moody and Gradv (2017) in
Detroit, MI). The RISK Screening Environmental Indicators-Geographic Microdata model, which models
transport and dispersion of air emissions using AERMOD, has been used to model chemical-specific
Toxics Release Inventory (TRI) releases and air Pb concentrations around point sources, based on what is
known about environmental fate and transport (e.g., Hill et al. (2021) in Syracuse, New York) (U.S. EPA.
2022).
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Biokinetic models have been developed at U.S. EPA that estimate levels of Pb in blood given
information on potential exposure to Pb in environmental media. The Integrated Exposure Uptake
Biokinetic (IEUBK) model was first created in the late 1980s and early 1990s to help evaluate Pb
exposure in children at potential Superfund sites. It was designed to allow users to predict whether BLLs
for children from birth to seven years over periods no less than a month exceed a target BLL based on a
GM BLL predicted from available information about exposure to Pb (SRC, 2020). The All-Ages Lead
Model (AALM) was developed to extend biokinetic modeling capability beyond the age of seven, include
intermittent exposures, and model additional (e.g., bone) tissue concentrations of Pb. Both of these
models are described in detail in Chapter 4 of the 2006 Pb AQCD (U.S. EPA, 2006), and recent updates
are described in Section 2.6 of this document. Sections 2.2, 2.3, and 2.4 discuss toxicokinetics, biomarker
measurements, and biomarker trends, respectively.
The Stochastic Human Exposure and Dose Simulation (SHEDS)-Multimedia model is a
U.S. EPA probabilistic model for estimating environmental exposures through inhalation, ingestion, and
dermal routes. Estimates of exposure are based on human activity recorded in the Consolidated Human
Activity Database, dietary consumption surveys, and modeled or observed levels of a contaminant in
environmental media, food, and surfaces. Zartarian et al. (2017) and Stanek et al. (2020) used the
SHEDS-Multimedia model in combination with an approximation of IEUBK to estimate drinking water
Pb contributions to blood Pb in U.S. children. The Zartarian et al. (2017) analyses included a comparison
of the coupled SHEDS-IEUBK methodology against CDC's national-scale NHANES BLL data and an
exposure pathway contribution analysis. Stanek et al. (2020) used the SHEDS-IEUBK methodology to
evaluate various drinking water scenarios" relationship to BLL. NORMTOX and Modeling Environment
for Total Risk are other models used for estimating environmental exposures to Pb, described in the 2013
Pb ISA (U.S. EPA. 2013).
2.1.3 Exposure Studies
The following Sections describe research on Pb exposure through various environmental media,
dietary sources, consumer products, and ammunition.
2.1.3.1 Airborne Pb Exposure
Airborne Pb exposure occurs through inhalation of Pb in air and can be measured most accurately
for an individual through personal air exposure monitoring. Although the 2006 Pb AQCD (U.S. EPA,
2006) contained limited data on personal exposure monitoring of airborne Pb, the 2013 Pb ISA (U.S.
EPA, 2013) expanded upon this issue. The 2013 Pb ISA (U.S. EPA, 2013) contains detailed information
on studies that show how outdoor, indoor, and personal Pb-PM concentrations were correlated and varied
by local conditions in studies ranging from 1999 to 2010 (Table 2-1 reproduced below).
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Table 2-1 Comparison of personal, indoor, and outdoor Pb-PM
measurements from several studies included in the 2013 Pb ISA
i Pb Sampling Personal di, Outdoor
study Location Metric Period Ph Indoor Pb p.
Clavton et al. (1999)
IL, IN, Ml,
Med.
July
13
6.6
8.5
MN, OH, Wl
Pb-PMso
1995-
(ng/m3)
May1997
Adqate et al. (2007)
Minneapolis-
Avg.
Spring,
6.2
3.4
2.0
St. Paul, MN
Pb-PM2.5
Summer,
(ng/m3)
Fall 1999
Molnar et al. (2007)
Stockholm,
Avg.
December
Homes: 3.4
Homes: 4.5
Sweden
Pb-PM2.5
2003-July
Schools:
Schools:
(ng/m3)
2004
2.5
4.6
Preschools:
Preschools:
1.8
2.6
Tovalin-Ahumada et al. (2007)
Mexico City,
Med.
April-May
26
56
Mexico
Pb-PM2.5
2002
(ng/m3)
Puebla,
Med.
April-May
4
4
Mexico
Pb-PM2.5
2002
(ng/m3)
Pekev et al. (2010)
Kocaeli,
Avg.
May-June
Summer:
Summer:
Turkey
Pb-PM2.5
2006,
34
47
(ng/m3)
December
Winter: 85
Winter: 72
2006-
January
2007
Avg.
May-June
Summer:
Summer:
Pb-PMio
2006,
57
78
(ng/m3)
December
Winter: 125
Winter: 159
2006-
January
2007
Rasmussen et al. (2007)
Windsor,
Med.
April 2004
311
124
221
Ontario,
Pb-PM2.5
Canada
(mg/kg)
Pb = lead; PM = particulate matter.
Studies evaluated in the 2013 Pb ISA (U.S. EPA, 2013) (Table 2-1) have shown personal Pb-PM
concentrations to be higher than indoor or outdoor concentrations. The National Human Exposure
Assessment Survey (NHEXAS) study (Clayton et al.. 1999) cited in the 2006 Pb AQCD (U.S. EPA,
2006), which sampled Pb in multiple exposure media across six states in U.S. EPA Region 5 in 1995—
1997, found personal air Pb concentrations to be significantly higher than indoor or outdoor Pb
concentrations. Adgatc et al. (2007) found average personal Pb-PlVb 5 concentrations to be roughly three
times higher than outdoor Pb-PlVb 5 concentrations and roughly two times higher than indoor Pb-PlVb 5
concentrations, in 1999. In contrast, a more recent study Stevens et al. (2014) found personal Pb-PM
concentrations to be lower than indoor or outdoor concentrations in a 2004-2007 survey.
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The Detroit Exposure and Aerosol Research Study measured personal, indoor, and outdoor PM2.5
mass components in six Detroit, Michigan neighborhoods over a 3-year period from 2004 to 2007 during
winter and summer (Stevens et al.. 2014). As mentioned in Williams et al. (2009). which contains details
of the design and implementation of the study, daily monitoring was performed from Tuesday to Sunday
and integrated over a 24-hr time period from 9:00 a.m. to 9:00 a.m. the next day. In contrast to previous
studies, which found higher personal Pb concentrations, the authors found personal Pb-PMo.s
concentrations were slightly less than indoor or outdoor concentrations at most sites. High relative
standard deviation (RSD) values associated with PM mass components in personal measurements (Pb
mass concentration RSD ranged from 66% to 270%, depending on the site) indicated high spatial and
temporal variability across sites, pointing to possible influence by personal activities and
microenvironments.
Table 2-2 shows outdoor, indoor, and personal Pb-PlVb 5 collected by season across six sites
(Stevens et al.. 2014). Sites were selected according to their proximity to suspected PM sources, including
industry (Sites 1, 4, and 5), diesel truck traffic (Site 3), automotive traffic (Sites 4 and 6) and results from
pilot testing of measurements of PM, carbon monoxide, and polycyclic aromatic hydrocarbon (PAH)
concentrations at those sites. Site 7 had very low concentrations of measured PM determined to be due
only to regional influences. Site 5 was in a heavily industrialized area and also showed the highest
concentrations of other elements measured in the study, including Fe, Mn, and Ca. Higher concentrations
of Pb in outdoor, indoor, and personal air measurements at Site 5 suggest that industrial emissions
contributed to Pb not only in outdoor air but also air that infiltrated homes and microenvironments.
Personal monitoring used active and passive monitors attached to a nylon vest, whereas indoor and
outdoor monitoring used similar monitors but with weather shielding. The overall Pb-PM2 5 mass
concentration ratio of personal to indoor air during summer and winter was 1.1 and 0.9, respectively. Our
literature search and screening did not capture other recent literature containing personal Pb-PM
concentration measurements. Section 2.5 explores BLLs and their relationship to Pb in air and in other
environmental media.
Table 2-2
Pb-PM2.5 concentrations across six sites in Detroit, Michigan
Site3
Season
Outdoor - Mean (ng/m3)
Indoor - Mean
(ng/m3)
Personal - Mean
(ng/m3)
1
Summer
14.0
12.0
11.0
Winter
9.0
6.0
6.0
3
Summer
9.0
8.0
8.0
Winter
8.0
5.0
6.0
4
Summer
4.0
4.0
4.0
Winter
5.0
3.0
3.0
5 Summer 15.0 12.0 11.0
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Site3
Season
Outdoor - Mean (ng/m3)
Indoor - Mean
(ng/m3)
Personal - Mean
(ng/m3)
Winter
42.0
14.0
13.0
6
Summer
6.0
4.0
5.0
Winter
5.0
2.9
3.0
7
Summer
5.0
2.0
4.0
Winter
4.0
3.0
3.0
aSite 2 was originally considered for inclusion; ultimately, however, its characteristics were deemed similar to some of those
already involved.
Data sou reed from Stevens et al. (2014).
Indoor air Pb concentrations can vary with outdoor air Pb concentrations because of infiltration
rates, indoor and outdoor Pb sources, and meteorology. As seen in Table 2-1 and Table 2-2 above,
although the majority of studies showed higher outdoor Pb-PM concentrations than indoor Pb-PM
concentrations, some studies have recorded higher indoor Pb-PM concentrations. Indoor dust containing
Pb may be disturbed and released into indoor air environments, contributing to indoor Pb-PM
concentrations. Resuspension rates due to foot traffic may be affected by walking behavior, type of floor
surface (e.g., carpet, vinyl), and particle size (U.S. EPA. 2013). Williamson et al. (2021) measured Pb in
PM at a high school in Texas by taking thirteen samples each integrated over 2-6 days for a period of
2 months. Elemental analysis showed an average Pb indoor-outdoor mass concentration ratio in PM10-2.5
to be 2.1, suggesting the presence of indoor sources.
The infiltration of outdoor Pb-PM can play a role in the relationship between indoor and outdoor
concentrations and is affected by multiple factors. A subsequent multivariate fixed effects analysis of the
NHEXAS-MD data (Clayton et al.. 1999) by Egeghv et al. (2005) found Pb levels measured in indoor air
were significantly associated with log-transformed outdoor air Pb levels, ambient temperature, number of
hours in which windows were open, whether the home was built before 1950, and fireplace usage
frequency. Molnar et al. (2007) measured PM2 5 in homes, preschools, and schools in Stockholm, Sweden
and found a net infiltration rate of -0.6. As shown in Table 2-2, Stevens et al. (2014) found overall ratios
of indoor to outdoor Pb during summer and winter to be 0.7 and 0.2, respectively, suggesting that outdoor
air had greater infiltration during summer.
Ambient Pb concentrations can vary spatially across urban centers because of point
(e.g., industrial facilities, airports) and nonpoint (e.g., roadway networks) sources, as well as the
meteorology (wind strength and direction) that disperses Pb. Section 1.5.3 of this ISA
(https://assessments.epa.gov/isa/document/&deid=359536) contains studies that examined spatial
variability of Pb air concentrations at several urban centers, such as Los Angeles, CA and St. Louis, MO
and were attributed to a wide variety of sources including nearby chat piles, abrasive vehicle emissions,
and a previously operating Pb smelting plant (Li and McDonald-Gillespie. 2020; Yadav and Turner.
2014; Pakbin et al.. 2011). Emissions from avgas can also contribute to Pb concentrations at and around
airports, as discussed in Section 1.2.1 of this ISA
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(https://asscssmcnts.cpa.go\ /isa/documcnt/&dcid=359536). These concentrations have the potential to be
inhaled by those on airport grounds or in surrounding neighborhoods. These emissions can also deposit
into soil surrounding airports or mix with suspended soil Pb concentrations in air. Some studies, discussed
in Section 2.4.1, have associated BLLs with proximity to airports that use avgas (Zahran et al.. 2017a;
Miranda et al.. 2011). Resuspension of Pb deposited from historical sources may also contribute to Pb
exposures. Section 2.4 discusses the relationship of BLLs to various Pb sources.
The size of Pb particles that someone may be exposed to can vary due to source type and
proximity to those sources. The size distributions of soil and house dust particles tend to be larger than
ambient air particles (Siciliano et al.. 2009; U.S. EPA. 1990; Hee et al.. 1985). Particles that are either
ultrafine or coarse may be affected by particle dynamics that limit their contribution to exposure. Before
ultrafine Pb-PM reaches a person, these particles may aggregate into larger sizes (Havs et al.. 2011). On
the other end of the size distribution spectrum, coarse particles have higher settling velocities than fine
and ultrafine particles, meaning that exposure to these larger particles will likely be more spatially and
temporally heterogeneous than fine particles, which can travel farther across urban centers (U.S. EPA.
2013). Studies examining the size distributions of Pb-PM are discussed in more detail in Section 1.5.5
(https://asscssmcnts.cpa.go\/isa/documcnt/&dcid=359536).
Cho et al. (2011) found that published literature after 1986 indicated a shift in the primary particle
size mode of airborne Pb particles from below 2.5 |im to between 2.5 and 10 |im. attributed to a shift
away from the use of Pb in motor vehicle gasoline. There is also evidence that Pb particles in emissions
from piston-engine aircraft are smaller than those emitted from an automobile engine using the same
leaded fuel. The addition of tetraethyl Pb in both avgas and motor vehicle gasoline results in exhaust
containing Pb dibromide particles. Previous studies of motor vehicle exhaust showed that these particles
range in size from around 20 to 100 nm in diameter with a mean of 50 nm (NASEM. 2021). Griffith
(2020) tested 100LL (100 octane, low Pb) avgas in a 1959 model aircraft and a 1957 model automobile.
Exhaust samples from the piston-engine aircraft were found to be 13 nm in average diameter, whereas
those from the automobile were 35 nm in average diameter. Both exhausts contained Pb dibromide beads
in a hydrocarbon matrix; however, the motor vehicle exhaust particles contained 5-10 beads or more,
whereas those in the aircraft exhaust were found to contain 1-2 beads.
2.1.3.2 Exposure to Pb in Soil and Dust
As described in detail in Section 1.3.2 (https://assessments.epa.gov/isa/document/
&deid=359536). Pb can be found in soil and dust as a result of deposition of atmospheric Pb, past
combustion of leaded gasoline, automobile parts (e.g., wheel weights), aviation, industrial activities, or
Pb-based paint. This Pb can be further transported through resuspension in dust back into ambient air or
tracked indoors and resuspended into indoor air environments. Pb in soil can contribute to exposure
through ingestion (i.e., hand-to-mouth activity) or inhalation of resuspended dust. Ingestion may also
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occur after mucociliary transport out of the ciliated airways into the esophagus, as described in
Section 2.2.1.1. As described in the following paragraphs, elevated Pb concentrations have been found in
a wide variety of outdoor soil locations, including residential properties, near roads, on or near airports,
playgrounds, urban gardens, and in house dust.
2.1.3.2.1 Outdoor Pb
Resuspended soil and dust can contribute to outdoor Pb-PM concentrations. Section 1.2.6 of this
current document (https://asscssments.epa.go\ /isa/documcnt/&dcid=359536) contains detailed
information on recent research investigating the contribution of resuspended soil to airborne
concentrations. Laidlaw et al. (2012) analyzed the contribution of resuspended soil Pb to air Pb in
Birmingham, AL; Chicago, IL; Detroit, MI; and Pittsburgh, PA. Using the IMPROVE soil estimation
calculation, which estimates soil content in the air based on the primary components of soil (Al, Si, Ca,
Fe, and Ti), and data from one sampling location in each city over different time periods, the authors
found atmospheric Pb strongly correlated with atmospheric soil concentrations. Using a mixed effects
model, the predicted percent increase in atmospheric Pb for each percent increase in atmospheric soil was
0.709 (95% CI: 0.535, 0.882) in Pittsburgh, 0.848 (95% CI: 0.724, 0.973) in Detroit, 0.710 (95% CI:
0.573, 0.847) in Chicago, and 0.922 (95% CI: 0.812, 1.033) in Birmingham.
Pb in soil has been found above background levels in both major urban centers and smaller cities
(Clark and Knudsen. 2013). As discussed in Section 2.2.1.2.3, Pb has been found to be consistently
enriched in soil particles <150 |im: however, Pb enrichment in this size range is usually not considered as
part of the soil sampling protocol in various studies. A study of 170 homes in Appleton, WI found a range
of Pb concentrations between 47 and 32,483 |ig Pb/g among soil around homes of various types. Soil next
to homes built before 1960 had significantly (p < 0.001) higher GM Pb concentrations than homes built
after 1960, and spatial sampling of soil next to a subset of 71 homes found a general decreasing trend in
Pb concentrations with increasing distance away from each home (Clark and Knudsen. 2013). Pavilonis et
al. (2020) measured Pb concentrations in soil from 34 parks across New York City; concentrations were
found to range from 7.8 mg Pb/kg to 6,300 mg Pb/kg, with a median concentration of 161 mg Pb/kg.
Parks in areas with the highest population growth between 2010 and 2017, greatest manufacturing
density, most new building construction, and greatest street density had the highest Pb concentrations;
however, how much these individual factors contributed was not resolved. Wang et al. (2022) captured 99
surface (2-3 cm of the mineral-soil surface excluding overlying organic materials) soil samples in
Durham, NC. Mean total concentrations (± standard deviation) of Pb were found to be 2,281 mg/kg
(±2,868 mg/kg, n = 31) in house foundation soils, 321 mg/kg (±533 mg/kg, n = 42) in urban streetside
soils, 42.1 mg/kg (±25.0 mg/kg, n = 19) in city park soils, and 15.9 mg/kg (±3.58 mg/kg, n = 7) in
suburban streetside soils. By using isotopic signatures, the authors found that house foundation soils had
significant input of legacy Pb-based paint, whereas urban streetside soils had a mixed origin made up
predominantly of legacy leaded gasoline and atmospheric deposition.
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The distribution of soil Pb concentrations in New Orleans has been particularly well documented.
Between 1989 and 2015, four surveys of soil Pb were carried out. For each survey, soil samples were
collected from street sides, near home foundations, and from vacant properties and parks as far as possible
from house sides and streets, and data were stratified by census tract. Due to the effects of Hurricanes
Katrina and Rita in 2005, Surveys 2 and 3 examined fewer census tracts than previous surveys but
stratified data according to the census boundaries established in Survey 2 for continuity. Because many of
the tracts not sampled in Surveys 3 and 4 were outlying areas with relatively low Pb concentrations,
comparison to the full data set from Survey 2 would not be appropriate. The results of these surveys are
summarized in Table 2-3, including average median Pb concentrations for the full data set from each
survey, as well as subsets of data from Survey 2 corresponding to only those census tracts included in
Surveys 3 and 4 for comparison. Key outcomes of these surveys are discussed below.
Table 2-3
Median soil Pb concentrations in New Orleans census tract level
surveys
Survey Number
Yr Number of Tracts
Number of Samples
Full Data Set Median
Cone, (mg/kg)
Subset Median
Cone, (mg/kg)
1
1989-1992
283
4,011
134
2
1998-2000
286
4,388
100
329a, 280b
3
2005-2006
46
1,748
203
4
2013-2015
176
3,320
132
yr = year(s).
aAverage median soil Pb concentration across the 46 census tracts included in Survey 3.
bAverage median soil Pb concentration across the 176 census tracts included in Survey 4.
Data sourced from Mielke et al. (2005). Zahran et al. (20101. and Mielke et al. (20161.
Mielke et al. (2005) first reported data from Surveys 1 and 2. In Survey 1, 71 of 286 census tracts
had median soil Pb that exceeded the U.S. EPA regulatory standard for certain residential properties of
400 mg/kg1, and 10 census tracts had a median soil level of >1,000 mg/kg. In general, Pb concentrations
in outlying suburbs decreased moderately from Survey 1 to Survey 2. However, the authors note this
decline was a result of moderate decreases in suburban soil Pb concentrations, and inner-city areas
actually increased. Mielke et al. (2011b) further examined a subset of samples from Survey 2. This study
includes 224 soil samples collected from ten Housing Authority (HA) public properties and 363 soil
samples collected from residential private properties within an 800 m radius of the centroids of the HA
properties. Six HA properties were located within the inner city, and the other four were located in the
1 As defined in the Code of Federal Regulations (40 CFR 745), a soil-Pb hazard is bare soil on residential real
property or on the property of a child-occupied facility that contains total Pb equal to or exceeding 400 ppm (|ig/g)
in a play area or an average of 1,200 ppm of bare soil in the rest of the yard based on soil samples.
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outlying areas of New Orleans. As observed in the full data set, samples retrieved from both HA
properties and private residences had higher Pb concentrations in inner-city locations compared with
those from outlying areas. However, Pb concentrations in soil taken from HA locations were significantly
lower (about half or less) than Pb concentrations taken from nearby private residences. The authors
attribute this difference to the fact that HA properties had unpainted brick facades, whereas private
properties were painted wood. Altogether, these results highlight the importance of legacy inputs of Pb
from leaded automobile gas and Pb paint on soil Pb concentrations.
Zahran et al. (2010) examined soil Pb concentrations from Survey 3 compared with soil Pb
concentrations in those census tracts from Survey 2. In 29 of 46 neighborhoods examined, median soil Pb
declined between surveys. In Survey 3, 6 of 46 census tracts had soil Pb levels of >400 mg/kg, compared
with 15 of 46 neighborhoods exceeding this standard in Survey 2. Mielke et al. (2016) compared Pb
concentrations in soil from Survey 4 to those measured in the same census tracts in Survey 2. Median soil
Pb levels across sampled census tracts dropped significantly, which the authors attribute to factors
associated with Hurricane Katrina rather than reduction in inputs. Specifically, they cite removal of Pb-
painted drywall and woodwork during renovations and sequestration of Pb-contaminated soil beneath
low-Pb sedimentary material from outside the city, which was moved in both intentionally during
reconstruction and unintentionally during levee breaches associated with the storm (Mielke et al.. 2000).
They theorized that addition of clean soil may provide an effective means of mitigating the impact of Pb
exposure. This method was further explored by Walsh et al. (2018) and Egendorf et al. (2018) and found
to be effective. However, the findings of Rabito et al. (2012) indicate that elevated Pb concentrations
persist in many areas of New Orleans. This study used a different sampling strategy than the census tract
studies, focusing solely on Pb concentrations in soil samples taken near homes. Of the 109 homes
sampled in 2009, Pb concentrations were often still high; nearly half had elevated soil Pb, and 27% of
those homes had soil Pb greater than 1,200 ppm.
Pb has also been found present in children's playground soils, which is a concern for accidental
ingestion by children (U.S. EPA. 2013). Mielke et al. (2011a) found Pb soil concentrations ranged from
14 to 3,692 mg/kg with a median soil concentration of 558 mg/kg on playground soils at 11 daycares and
community centers in New Orleans. Almansour et al. (2019) found, among 28 randomly sampled
playgrounds in Boston, a median Pb soil concentration of 65.7 mg/kg. This number was typical for soils
in Massachusetts based on U.S. Geological Survey data, indicating this was likely background Pb instead
of Pb originating from anthropogenic sources (Smith et al.. 2013).
Human uptake of Pb through soil exposure can also occur during gardening as a result of
unintentional soil and dust ingestion. Gardeners may not wear protective equipment or properly wash off
soil when finished with gardening, and they may eat or drink while working or track soil into the home
(Schmeltz et al.. 2020; Spliethoff et al.. 2016). However, exposure to soil Pb in urban gardens can be
lower than in other gardens as a result of clean soil being brought in for gardening beds (Spliethoff et al..
2016). The 2013 Pb ISA (U.S. EPA. 2013) evaluated studies that previously looked at Pb in urban garden
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soil, including Clark et al. (2006). who tested the soil in 103 urban gardens in two Boston neighborhoods.
Using isotopic analysis, the authors found that Pb-based paint contributed 40 to 80% of Pb in the urban
garden soil samples, with the rest coming from historical Pb emissions. Furthermore, Clark et al. (2006)
estimated that Pb consumption from urban gardens could be equivalent to 10 to 25% of the exposure to
Pb from drinking water for children living in the Boston neighborhoods studied. Spliethoff et al. (2016)
investigated bed soil (508 samples, where plants were being grown) and non-bed soil (54 samples, where
no growing was occurring) of urban gardens in New York City and found bed soil had a median of 96 mg
Pb/kg, whereas non-bed soil had a median of 181 mg Pb/kg. The authors also estimated mean dust-Pb
concentration due to soil tracking at 72 mg Pb/kg. In a separate study, Cheng et al. (2015) found a median
value of 355 mg Pb/kg in 1,652 urban garden soil samples around New York City. Three percent of
community garden samples and 18% of home garden samples were found to exceed 1,200 mg Pb/kg, a
level of contamination not recommended for vegetable gardening (U.S. EPA. 2014). A small pilot study
of urban gardens in New York City found a mean value of 372 ppm Pb in the 18 soil samples taken
(Schmeltz et al.. 2020).
In urban gardens, there can be spatial variability across Pb concentrations at the surface. A study
in Terre Haute, IN collected 1,061 surface soil samples from a 1.25-acre (54,450-sq. ft.) urban garden at
high spatial resolution. All samples were collected from the top several inches of soil and stored in sample
bags for analysis using handheld XRF. The authors found there was high variability across the garden,
ranging from background Pb levels to concentrations above 800 ppm (Latimer et al.. 2016). A smaller
study of an urban garden in southern Detroit, MI, which collected 80 samples, found a mean value of
151 mg Pb/kg among all samples. However, there was also high variability across samples taken, with a
minimum of 17 mg Pb/kg and a maximum of 882 mg Pb/kg found (Bugdalski et al.. 2014).
Soil and dust transfer can be an important exposure route for Pb. This is especially true for
children who may play outside, close to the ground (Mova and Phillips. 2014). The updated U.S. EPA
Exposure Factors Handbook (U.S. EPA. 2017) reviewed soil/dust ingestion studies and based on that
literature recommends, for use in general population modeling or risk assessment, a daily soil and dust
ingestion rate made up of a combination of soil and settled dust of 40 mg/day (<6 months old), 70 mg/day
(6 months to <1 year), 90 mg/day (1 to <2 years), 60 mg/day (2 to <6 years), 80 mg/day (1 to <6 years),
60 mg/day (6 to <12 years), and 30 mg/day (12 years through adult) (U.S. EPA. 2017). These are central
tendency values for the general population. Ingestion rates through hand-to-mouth transfer can be
important for determining exposure to soil and dust particles. In a review of studies examining soil and
dust ingestion rates in children, Mova and Phillips (2014) noted mean daily ingestion rates varied by
quantification method, each of which has specific limitations. Studies that used the tracer element
method, biokinetic model comparison method, and activity pattern method reported mean daily soil and
dust ingestion rates of 26-470, 110, and 10-1,000 mg/day, respectively, von Lindern et al. (2016)
estimated soil and dust ingestion rates for multiple age groups of children (<10 years of age) by using
BLL data, age-specific biokinetic slope factors, and estimated Pb uptake. Mean ingestion rates ranged
from 50 to 154 mg/day across all age groups and scenarios. The central tendency soil and dust ingestion
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rates for children from von Lindern et al. (2016) are the recommended defaults in IEUBK model version
2.0. An evaluation of the IEUBK model version 2.0 using ingestion rates from either von Lindern et al.
(2016) or (U.S. EPA, 2017) showed that children's predicted BLLs were well aligned with observed
BLLs (Brown et al.. 2022; U.S. EPA. 2021a).
Two recent studies have used the Stochastic Human Exposure and Dose Simulation Soil and Dust
(SHEDS-Soil/Dust) model to estimate distributions of soil and dust ingestion rates in children and adults.
Ozkaynak et al. (2022) reported arithmetic mean ingestion rates (dust plus soil) of approximately
40 mg/day for children <1 year of age, 50 mg/day for 1 to <3 years, 60 mg/day for 3 to <11 years,
40 mg/day for 11 to <16 years, and 20 mg/day for adolescents 16 to <21 years. A notable contribution of
this work was the consideration of the season of the year and the time individuals spent outside. For
children <1 year of age, ingestion was from dust only due to an assumed negligible amount of time
outside. For other age groups, daily soil-only ingestion rates were approximately doubled (based on
Tables S7-S10 of the paper) in the summer relative to other seasons, increasing from 9 to 15 mg/day for
children 1 to <2 years of age, 20 to 47 mg/day for 2 to <3 years, 23 to 57 mg/day for 3 to <11 years , 17 to
40 mg/day for 11 to <16 years, and 9 to 20 mg/day for 16 to <21 years. Hubbard et al. (2022) reported
that for adults (>21 years), daily average ingestion rates of soil and dust could range from 7 to
123 mg/day for the general population to high occupational exposures, respectively. This study also
showed soil ingestion rates were increased in the summer relative to other seasons. Dust ingestion rates
were only minimally affected by seasonality.
The properties of soil, including size and humidity, can affect the adherence of soil particles to
hands. Ruby and Lowney (2012) suggests particle adherence occurs below 150 |im. Finer soil particles
(<63 (mi in diameter) tend to adhere to human hands more efficiently than larger particles. Soil with
higher moisture content results in slightly larger particles (<100 |im in diameter) having selective
adherence to hands (U.S. EPA, 2017). Approximately 90% of the cumulative mass of soil adhered to
children's hands is <150 |im in size. Smaller particles are more mobile than larger particles and are more
likely to accumulate in the indoor environment as a result of deposition of wind-blown soil or track-in
transport of soil on clothes, shoes, pets, toys, and other objects, providing additional opportunity for
exposure to this particle size fraction (Stalcup, 2016).
2.1.3.2.2 Indoor Pb
Both the 2006 Pb AQCD (U.S. EPA. 2006) and the 2013 Pb ISA (U.S. EPA. 2013) recognize
house dust as a pathway for Pb exposure. Table 3-5 of the 2013 Pb ISA (U.S. EPA. 2013) contains studies
that measured indoor Pb dust concentrations between 2006 and 2011. Median Pb dust concentrations
ranged from 63 mg/kg (Zotaet al.. 2011) to 470 mg/kg (Spalinger et al., 2007), although locations
sampled within buildings (e.g., floors, windowsills) and sampling procedures varied by study.
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The 2013 Pb ISA (U.S. EPA. 2013) discusses how Pb in house dust can be present as a result of
infiltration from outdoors. Pb-containing dust and soil may enter a building through infiltration in the air
or on the surfaces of objects and persons who enter the building. Proximity to historic and active metals
mining and smelting sources has been linked to increased levels of Pb in house dust (Zota et al.. 2011;
Gaitens et al.. 2009; Spalinger et al.. 2007). Tu et al. (2020) estimated contributions of yard soil to indoor
dust by collecting indoor residential dust and soil surrounding homes in eight communities near former
mining or smelting operations. Mass soil-to-dust transfer coefficients with good to moderate fit were
found to range from 0.14 to 0.47 for Pb.
Pb can be released from housing materials, often linked to older homes that may have Pb-based
paint (Mielke and Gonzales. 2008). Dietrich et al. (2022) found that among 434 sampled homes around
the United States, exterior paint peeling, interior paint peeling, and older housing were predictors of
higher Pb dust concentrations. Pb-containing dust can be present in carpet and on other flooring material
(Wilson et al.. 2007; Yu et al.. 2006). The second American Healthy Homes Survey (AHHS II) concluded
in June 2019 and sampled 703 homes in 37 states for Pb-based hazards. Results from the AHHS II
estimate that 34.6 million homes have Pb-based paint somewhere in the building, a decrease from both
AHHS I, which estimated 37.1 million homes, and the National Survey of Lead and Allergens in
Housing, which estimated 37.9 million homes. The AHHS II also estimates that 29.0 million homes had a
dust Pb hazard present, defined as a dust Pb level >10 |ig/ft2 or a windowsill dust Pb level >100 (ig/ft2
(U.S. EPA. 2020). Sowers et al. (2021) used X-ray absorption spectroscopy on a small subset of dust and
soil samples collected from homes used in the AHHS I and found that Pb-based paint contributed strongly
to house dust. A study of 102 homes in Rochester, NY with unenclosed painted porches found that 92%
tested positive for Pb-based paint (>1 mg/cm2). The GM on the tested components was 1.1 mg/cm2 (95%
CI: 0.88 mg/cm2, 1.141 mg/cm2) (Wilson et al.. 2015).
In another study, over 100 homes in Philadelphia, many within 0.5 miles of a legacy Pb point
source (industrial facilities, Pb-based paint), were found in 2014 to have median front door floor dust,
mean child play area floor dust, and child bedroom windowsill dust Pb levels of 17.7 (ig/ft2, 13.9 |ig/ft2.
and 31.2 (ig/ft2, respectively (Dignam et al.. 2019). A small study of 35 homes in the United States found
that Pb concentrations in floor varnish were correlated with pre-1930s housing during reflnishing
exercises (Schirmer et al.. 2012). Matt et al. (2021) investigated the contribution of tobacco smoke to Pb
dust concentrations in the homes of 60 multiunit housing residents in San Diego, using wipe and vacuum
floor dust samples. Vacuum dust nicotine loading was found to be significantly (p = 0.0012) associated
with Pb dust loading; however, vacuum dust nicotine was not found to be associated with Pb
concentrations in surface wipes from floors or windows. Floor wipe samples of nicotine concentrations
were also not associated with Pb measured in vacuum dust or surface wipe samples.
Dust-Pb concentration values are important for calculating estimates of Pb intake or as input to
blood Pb models (e.g., IEUBK), whereas dust-Pb loading values can be compared with dust-Pb loading
regulatory values and can serve as another representation of exposure (Bevington et al.. 2021). However,
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of the many dust-Pb monitoring studies that exist, only some report both dust-Pb concentration and dust-
Pb loading values (i.e., the dust-Pb concentration multiplied by the total dust loading on a surface).
U.S. EPA previously combined data from three studies to create a dust-Pb loading to dust-Pb
concentration (LTC) model based on empirical data (U.S. EPA. 2019d). Bevington et al. (2021)
developed an LTC model by pairing 2,174 dust-Pb loading and dust-Pb concentration values across five
studies (each with n > 200 homes), incorporating data from an additional two studies not used in the EPA
model (Clayton et al.. 1999; Lanphear et al.. 1996). The authors evaluated 17 different versions of the
LTC model across a wide range of dust-Pb loadings (0.1-10,000 (ig/ft2) using evaluation data from 32
studies and found there was relatively good agreement between the LTC models and data sets for central
tendency values of dust-Pb concentrations. The model with the most agreement, model 16, had slope
(0.413), y-intercept (5.291), and R2 (0.578) values that were overall most similar to the evaluation dataset
slope (0.440), y-intercept (5.511), and R2 (0.473) values among the different LTC models tested. At high-
dust Pb loadings, the predicted values for dust-Pb concentrations were overestimated; however, the
highest dust-Pb loading values came from intervention studies in which dust-Pb loading was likely higher
than would occur in homes found in the general population.
2.1.3.3 Dietary
Possible sources of Pb in food include introduction during processing or preparation with Pb-
contaminated drinking water, preparation in Pb-glazed cookware, deposition of Pb onto raw food
materials, uptake from soil by fruit and vegetable crops, and Pb exposure in livestock that produce dairy
or meat ingredients (U.S. EPA, 2013). Pb in commonly consumed food items purchased from grocery
stores in the United States is measured and reported on an ongoing basis by the Food and Drug
Administration (FDA) Total Diet Study (TDS). These data have been combined with food consumption
data from What We Eat in America (WWEIA), the food consumption section of NHANES, to model
dietary Pb intake (Gavelek et al„ 2020; Spungen. 2019). Because of the high risk of Pb poisoning
associated with low body mass, dietary Pb in infants and children is a particular concern and the FDA has
issued updated interim reference levels (IRLs) for dietary Pb intake of 2.2 (ig/day for children and
8.8 (ig/day for women of childbearing age (Flannery and Middleton, 2022).
The 2006 Pb AQCD (U.S. EPA. 2006) stated that according to the TDS data for surveys
conducted between 1982-1984 and 1994-1996, estimates of Pb intake from food dropped across all age
groups (U.S. EPA, 2006). This was attributed to a general decline in food Pb concentrations resulting
from regulations, such as a ban on Pb soldering in food cans and the ban on Pb additives in automobile
gasoline, which reduced contamination in crops and livestock. However, the 2013 Pb ISA (U.S. EPA,
2013) summarized results of the 2008 TDS, which found a range of Pb concentrations in foods, with the
highest levels (>65 |ig/kg) measured in noodles, carrots in baby food, and oatmeal in baby food (U.S.
EPA, 2013). The document also summarized the results of Manton et al. (2005). which suggested that
some dietary Pb in children 0-12 months may originate from Ca salts used in some baby formula. These
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findings demonstrate that although concentrations of Pb in food have generally fallen, there is
considerable variability, underscoring the importance of considering individual behavior when assessing
risk associated with dietary Pb exposure.
Zartarian et al. (2017) investigated dietary Pb exposure based on 2007-2013 TDS data and found
diet may still be a major contributor to BLLs for some individuals. The study used a combined SHEDS-
IEUBK multimedia model and data for children's activity patterns, Pb concentrations in media, exposure
factors, and biokinetic dose factors from a variety of sources that were intended to simulate exposure
conditions for 2009-2014 NHANES data and NHEXAS Region 5 data. The authors found ingestion of
soil/dust, food, and water were major contributors to BLLs. However, results varied depending on the age
of the participant and BLL percentile. For 1 to <2-year-olds soil/dust ingestion was the dominant pathway
above the 80th percentile, but food intake was a major contributor below the 70th percentile, accounting
for half of blood Pb values and contributing -0.6 (ig/dL on average across all percentiles. Water
contributed ~10%—15% of the BLL, depending on the percentile, or -0.2 (ig/dL on average.
In addition, two studies used Pb concentrations reported in the 2014-2016 TDS surveys with
2009-2014 WWEIA food consumption data to model potential dietary Pb intake for specific groups.
Gavelek et al. (2020) estimated dietary Pb exposure in male and female children 7-17 years (n = 4,906),
women of childbearing age 16-49 years (n = 4,562), and men and women 18+ years (n = 14,614).
Spungen (2019) estimated dietary exposures for male and female children 1-6 years of age (n = 3,103)
and two subgroups of ages 1-3 (n = 1,717) and 4-6 years (n = 1,386). In both studies, lower-bound Pb
concentrations were calculated by setting all Pb values less than the limit of detection (LOD) to zero, and
upper-bound Pb concentrations were calculated by setting all Pb values less than the LOD to the LOD. In
addition, "hybrid" mean Pb concentrations were calculated by setting Pb values less than the LOD to zero
if there were no detected levels of Pb in food from 2009 to 2016; if Pb was detected in food at least once
from 2009 to 2016, Pb values less than the LOD were set to half the current LOD. Table 2-4 below shows
estimated mean and 90th percentile dietary Pb exposures for each population considered. For all groups
with children, the upper bound values on the mean and 90th percentile estimates for dietary Pb intake
exceed the current IRL for dietary Pb intake. In addition, the "hybrid" and lower-bound values on the
90th percentile estimates also exceed the current IRL. Notably, most food groups with the highest
contributions were related to highest consumption rather than highest Pb concentration in those foods.
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Table 2-4 Dietary exposures to Pb based on U.S. Food and Drug
Administration Total Diet Study (2014-2016) and What We Eat in
America (2009-2014) food consumption data
Dietary Pb Exposure (pg/day)
Mean 90th Percentile
Reference
Population
Lower
Bound3
Upper
Boundb
Hybrid0
Lower
Bound3
Upper
Boundb
Hybrid0
Male and Female 7-17 yr
1.4
4
2.2
2.3
5.8
3.4
Gavelek et al.
(2020)
Female 16-49 yr
1.6
4.6
2.4
2.8
6.7
4
Male and Female 18+ yr
1.7
5.3
2.7
3.2
7.8
4.5
1-6 yr
1.2
3.2
1.8
2
4.6
2.9
SDunaen (2019)
1-3 yr
1.0
3.0
1.7
1.8
4.4
2.6
4-6 yr
1.3
3.4
2.0
2.1
4.8
3.1
FDA = U.S. Food and Drug Administration; yr = year(s).
aValues less than LOD set to zero.
bValues less than LOD set to LOD.
°Values less than LOD set to zero if there were no detections from 2009 to 2016; otherwise, values less than LOD set to 0.5 * LOD.
Both Gavelek et al. (2020) and Spungen (2019) limited the scope of their analyses to data
collected from 2014 onward because FDA started using ICP-MS in 2014 to measure elemental
concentrations, as opposed to AAS, which was used previously. ICP-MS has a lower LOD and limit of
quantification than AAS (Gray and Cunningham. 2019). making these measurements difficult to compare
directly to past TDS results. Despite this change, 74% of samples measured in the 1994-1996 TDS were
below the detection limit for Pb, whereas in the new 2018-2020 TDS, this value increased to 85% of
samples, further demonstrating an overall decrease in food Pb concentrations. In addition, Spungen
(2019) found there may have been some decline in children's lower-bound mean Pb exposure from 2004-
2008 to 2014-2016, with 0.11 (ig/kg bw/day for 2-year-olds changing to 0.08 |ig/kg bw/day for 1 to 3-
year-olds, respectively.
While FDA TDS is a valuable tool for monitoring Pb in foods, the study design does have
inherent limitations. First, because sampling for this program is meant to be broadly representative rather
than comprehensive, more detailed studies may be valuable to fully describe Pb concentrations in food
items consumed primarily by children and women of childbearing age. For instance, a study carried out
by Gardener et al. (2019) presented data on Pb concentrations from an extensive sampling of baby foods.
Of the 564 U.S. baby food samples, Pb was detected in 37% of samples (median = non-detect,
max = 183.6 |ig/kg). but none exceeded FDA consumption guidelines. In addition, because the TDS
focuses on commonly consumed food items purchased at supermarkets, dietary Pb exposure from foods
not purchased from supermarkets may be overlooked. Small farms, home agriculture, and game meats
also present Pb exposure risk from dietary sources not captured in the TDS. Consumption of game meat
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hunted with Pb ammunition may increase dietary Pb exposure risk due to inadvertent consumption of
ammunition fragments as previously described in detail in the 2013 Pb ISA (U.S. EPA. 2013).
Spliethoff et al. (2014) investigated Pb in eggs of chickens raised in New York City community
gardens. Median Pb concentrations were found to be below the detection limit of 10 (.ig/kg. less than Pb
found in chicken eggs in previous studies (Van Overmeire et al.. 2009; Trampel et al.. 2003). Leibler et al.
(2018) investigated Pb in backyard-produced chicken eggs and modeled their contribution to children's
(younger than 7 years) BLLs using the IEUBK model. They found Pb egg concentrations were correlated
with surrounding soil amounts. Contributions to BLLs based on the IEUBK tested over four different
scenarios and different simulated ages ranged from 0.1 (ig/dL to 1.5 (ig/dL, depending on the frequency of
consumption and age of the child. In addition, Lupolt et al. (2021) measured Pb concentrations in 13
commonly consumed produce items sampled from 104 urban farm and community garden sites in
Baltimore, MD. Pb concentrations (ppb) measured in collards (58.3 ± 48.3), kale (58.3 ± 32.5), lettuce
(68.0 ± 121.0), cucumbers (23.6 ± 26.0), and peppers (51.7 ± 49.4) grown in the urban farms and gardens
were significantly higher (p < 0.05) than concentrations measured in store-bought, commercially grown
produce of the same type.
The 2013 Pb ISA (U.S. EPA. 2013) previously reported on Pb uptake and bioaccumulation in
agriculture. Uptake of Pb has been shown to occur in potted plants (Del Rio-Celestino et al.. 2006).
vegetable crops (Lima et al.. 2009). grasses (Vandenhove et al.. 2009). and wild mushrooms (Sesli et al..
2008). Pb contamination can occur from atmospheric deposition (Uzu et al.. 2010) and from treatment of
crops with compost produced from wastewater sludge (Cai et al.. 2007) and from fertilizer (Chen et al..
2008). Egendorf et al. (2021b) investigated the relative importance of Pb accumulation through roots,
splash, and atmospheric deposition in lettuce grown in soil with high (-1,200 mg/kg) and low
(-90 mg/kg) Pb concentrations in New York City and Ithaca, NY. In low-Pb soils, splash and
atmospheric deposition accounted for 84% and 78% of lettuce Pb grown in New York City and Ithaca,
respectively. In high-Pb soils, splash and atmospheric deposition accounted for 88% and 93% of Pb in
lettuces, with splash being the dominant mechanism. The authors also show soil covers, such as mulch,
were significantly (p < 0.05) correlated with lower Pb concentrations in lettuce compared with the bare
soil treatment due to reduced contamination from splash. Pb accumulation in agricultural crops is covered
in greater detail in Appendix 11.
2.1.3.3.1 Drinking Water
Drinking tap water is a pathway to Pb exposure. Several recent studies have been conducted that
focus on broad scale surveys of Pb concentrations in drinking water across a region and provide insight
on the subject. Sansom et al. (2019) looked at the exposure to Pb-contaminated drinking water in 13
residences of a Houston ship channel community. Pb concentrations above detection limits were found in
4 of the 13 homes studied, ranging from 0.6-2.4 ppb. Gleason et al. (2019) analyzed water systems used
in New Jersey in two distinct time periods: 2000-2004 and 2010-2014. Among all the water systems
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analyzed for Pb, 443,936 had Pb concentrations between 0 |ig/L and 2 |ig/L. and 7,845 had Pb
concentrations over 2 (ig/L. Desimone et al. (2020) analyzed 500 tap water samples from 72 Tennessee
schools in 2017 and 3,428 samples from 160 Tennessee schools in 2019. Pb concentrations detected
across all samples ranged from <0.5 (elementary schools, 2017) to 18,800 g/L (elementary schools,
2019), with medians ranging from 3.0 (middle schools, 2017) to 227 |ig/L (elementary schools, 2019)
among all samples. Nearly 90% of the schools tested (n = 205) had a Pb concentration higher than 1 |ig/L:
50 schools had a Pb concentration higher than 15 |ig/L. The average ages of the elementary, middle, and
high school buildings were >50 years old.
One of the primary factors driving the observed variability in drinking water Pb concentrations is
the corrosion of Pb plumbing components found in some homes and/or distribution systems; this can
include Pb service lines (LSLs), Pb-soldered joints, and Pb brass faucets and fixtures. Although new
LSLs were banned in 1986, drinking water infrastructure built prior to 1986 may still contain Pb
components, putting people residing in older homes and communities at greater risk. A 2016 survey of
infrastructure found that between 15 and 22 million people in the United States are served by community
water systems with full or partial LSLs (Cornwell et al.. 2016). This represents a significant reduction
from earlier surveys because of efforts to replace LSLs with safer alternatives. However, jurisdictional
issues sometimes make full replacement of LSLs impossible, resulting in implementation of a partial
replacement strategy in some areas. Trueman et al. (2016) evaluated the effects of full and partial
replacement of LSLs on Pb concentrations in drinking water in 45 single family homes. Prior to line
replacement, 90th percentile Pb concentrations in the first four L of water collected, beginning with the
first draw following a minimum 6-hour standing period, ranged from 16.4 to 44 |ig/L. For homes with full
replacement of LSLs, 1 month after replacement, 90th percentile Pb concentrations in tap water samples
ranged from 2 to 12 |ig/L. On the other hand, Pb concentrations in tap water collected from homes with
partial replacement of LSLs increased substantially in the first month and did not show a significant
reduction in concentration over the 6-month period of study. This is attributed to galvanic corrosion at the
interface between new copper plumbing and existing Pb pipes, which increases release of Pb to drinking
water. Similar results have been observed for BLLs associated with partial replacement of LSLs. Brown
et al. (2011) conducted cross-sectional analyses to determine whether children residing in houses with
LSL or partial replacement of LSL in Washington, DC had higher BLLs compared with children residing
in houses with no LSLs in Washington, DC. Between 2004 and 2006, children living in houses with
partially replaced LSLs were more likely to have a higher BLL compared with children living in houses
with no LSLs (OR = 1.9 [95% CI: 1.5, 2.3] for a BLL between 5 and 9 ^ig/dL; OR = 3.3 [95% CI: 2.2,
4.9] for a BLL >10 (ig/dL; relative to a BLL of <5 (ig/dL).
Many factors control the degree of corrosion in plumbing components and resulting Pb
mobilization to tap water. These include water treatment chemicals, pH, types and amounts of minerals
found in the water, age of Pb plumbing components, and water temperature. Corrosion control in LSLs
often involves developing an insoluble scale of Pb minerals that limits mobilization to water. This process
is facilitated by low temperature, high pH conditions with significant chlorine residuals from disinfection.
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Orthophosphate-based corrosion control inhibitors may also be added to sequester Pb in a less soluble
mineral phase. In recent years, water treatment plants in many municipalities have discontinued the use of
chlorine for disinfection due to the formation of carcinogenic byproducts. Commonly, chloramines are
used instead but may lead to greater mobilization of Pb to water due to the formation of more soluble Pb
minerals at a neutral pH (Rcnncr. 2006; Yasquez et al.. 2006). Because water quality depends on many
factors, U.S. EPA recommends new water treatment systems be optimized for corrosion control and any
subsequent change in treatment or raw water quality be assessed for potential increases in Pb
concentrations (U.S. EPA. 2003c).
Gibson et al. (2020) analyzed and merged the BLLs of 59,483 children in North Carolina with
demographic data and drinking water source (private wells or regulated water utility). The authors found
that among the children (n = 7,709) who drank from private wells, there was an increased chance of
higher BLL (mean =1.75 (ig/dL for private wells versus 1.59 (ig/dL for water utilities). Adjusting for all
other variables, the odds of an EBLL (defined as >5 (ig/dL in this study) was 2.1% for private well
drinkers versus 1.7% for water utility drinkers. These findings suggest that Pb released through corrosion
in private well systems may lead to increased BLLs as private wells are not covered under the Safe
Drinking Water Act and owners of private wells may not be using proper corrosion control (Knobeloch et
al.. 2013). Past studies have found higher levels of Pb concentrations in private well water (Stillo and
MacDonald Gibson. 2018; Pieper et al.. 2015).
Pb contamination in drinking water due to not implementing correct corrosion control methods
occurred in Flint, MI. Between 1967 and 2014, the city purchased treated water from the Detroit Water
and Sewage Department (DWSD), originating from Lake Huron. During this period, the Flint Water
Service Center (FWSC) was maintained as a backup treatment facility, treating water from the Flint River
only two to four times a year for a few days at a time and then discarding the treated water. However, in
2014, city officials made the decision to stop purchasing DWSD water and instead distribute water from
the Flint River with treatment at the FWSC. The DWSD water, originating from Lake Huron, was
optimized for corrosion control and treated with phosphate corrosion control inhibitors. However, the
FWSC was not optimized; the facility had difficulty maintaining chlorine residuals throughout the
distributions system for some periods and did not add corrosion inhibitors, both of which likely
contributed to destabilization of Pb scales (Masten et al.. 2016). In addition, because of a switch from
sulfate to chloride-based coagulants, the chloride to sulfate mass ratio (CSMR) increased from 0.45 to
2.04 (Pieper et al.. 2017). For water with alkalinity observed at the Flint facility (<50 mg/L), a CSMR
over 0.5 indicates very high risk for corrosion (Masten et al.. 2016). Water from the Flint River was
treated and distributed from April 2014 to October 2015. Independent studies found high concentrations
of Pb in drinking water taken from homes. Pieper et al. (2017) tested a home termed "Ground Zero" and
found Pb concentrations in water were well above actionable limits, ranging from 217 to 13,200 |ig/L in
April 2015. The same group carried out a larger study, examining Pb concentrations in tap water from
2015 to 2017 (Pieper et al.. 2018). In August 2015, the median Pb concentration in first draw water
samples from the 156 homes sampled was 3.5 |ig/L. with 17% of samples exceeding 15 |ig/L and a
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maximum measured concentration of 158 |ig/L. Samples taken after the reintroduction of DWSD-treated
water had lower median Pb concentrations in first draw water samples, at 1.9 |ig/L in March 2016 and
1.2 |ig/L in July 2016. However, while Pb concentrations in most homes decreased after reintroduction of
DWSD water, some remained anomalously high, either because of scouring of loose Pb deposits from
pipes or galvanic corrosion of partially replaced LSLs. Pb concentrations in drinking water for these
homes did not fall below 10 |ig/L until full replacement of LSLs (Mantha et al., 2020).
It is important to note between-study variation in the sampling and analytic methodologies may
reduce the comparability of Pb concentrations in tap water across studies. Factors including water sample
volume collected, stagnation, spacing of Pb within piping, and sampling protocol can all affect
measurements of Pb within tap water (Triantafyllidou et al„ 2021). Riblet et al. (2019) compared different
sampling protocols including different rates of flushing and stagnation after flushing in 21 households.
The authors found Pb concentrations in tap water ranged from 5.5 to 14.0 |ig/L. depending on the
sampling protocol used.
2.1.3.3.2 Breast Milk
Breast milk has been identified in prior reviews as a potential dietary source of Pb exposure for
infants (U.S. EPA, 2013, 2006). Table 2-5 shows the contribution of maternal blood Pb to Pb in breast
milk over the past 40 years in populations in the United States, Mexico, and Europe, as estimated from
data reported in papers. Gulson et al. (1998a) cautioned that studies reporting >1.5 |ig/L milk Pb per
(ig/dL blood Pb are likely due to contamination of samples (e.g., contamination due to Pb on hands of
women as they collect their milk). Trends in breast milk Pb concentrations with time postpartum
generally show no temporal relationship or a slight decline (Ettinger et al., 2006; Sowers et al., 2002;
Gulson et al., 1998a). Statistically significant Spearman correlations between breast milk Pb and both
whole blood Pb (r = 0.44, n = 81) and plasma (r = 0.31, n = 81) have been observed (Ettinger et al„ 2014).
In a study of healthy infants (97 males, 113 females; median age: 11.4 months; range: 8-23 months)
conducted from July 2014 to June 2016 in Seoul, Korea, duration of breastfeeding was correlated
(r = 0.427, p < 0.001) with the infants" BLLs (Choi et al„ 2017). Breastfed infants had significantly
(p < 0.001) elevated blood Pb (median: 1.12 (ig/dL; interquartile range: 0.77, 1.63) compared with mixed
fed infants" blood Pb (median: 0.81 (ig/dL; interquartile range: 0.51, 1.11) and formula fed infants" blood
Pb (median: 0.62 (ig/dL; interquartile range: 0.39, 0.82). Several recent studies and reviews have
investigated the effect of maternal Pb exposures and risk factors on breast milk Pb (e.g., Rebelo and
Caldas (2016) and Cherkani-Hassani et al. (2019)). These factors likely predominately reflect effects on
maternal blood Pb and are not further considered here.
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Table 2-5
Contribution of maternal blood Pb to breast milk at 1-3 months
postpartum
Milk Pb/
Blood Pb
Milk Pba
(M9/L)
Blood Pba
(Hg/dL)
nb
Location
Sample Yr
Reference
0.10
0.8 ± 0.7
7.7 ± 4.0
81
Mexico City, Mexico
1997-1999
Ettinqer et al. (2014)
0.15
1.4 ± 1.1
9.3 ± 4.5
310, 367c
Mexico City, Mexico
1994-1995
Ettinqer et al. (2006)
0.16
1.5 ± 1.2
9.4 ± 4.5
255
Mexico City, Mexico
1994-1995
Ettinqer et al. (2004a)
0.16
0.5 ± 0.3
3.1 ± 0.7
35
Holmsund, Sweden
1990-1992
Hallen et al. (1995)
0.24
2.8 ± 1.6
11.9 ± 9.4
39, 62c
Tucson, Arizona
yr not reported
Rockwav et al. (1984)
0.25
0.73 ± 0.70
2.9 ± 0.8
9
Migrated to Australia
yr not reported
Gulson et al. (1998a)
0.28
0.9 ± 0.4
3.2 ± 1.0
39
Roannskar, Sweden
1990-1992
Hallen et al. (1995)
1.35
1.74 ± 11.5
1.29 ± 0.60
cn
cn
o
Szczecin, Poland
2007-2008
Baranowska-Bosiacka et
al. (2016)
3.3
4 (median)
1 (median)
80
West Bank, Palestine
2017-2018
Shawahna (2021)
4.4
6.1 ± 1.0
1.4 ± 0.2
15
Camden, New Jersey
1997-2000
Sowers et al. (2002)
yr = year(s).
aMean ± Standard Deviation unless otherwise reported.
bPaired milk Pb and blood Pb samples unless otherwise indicated.
°Number of milk Pb samples, number of blood Pb samples, which includes additional subjects.
Ettinger et al. (2014) reported infant BLLs at 3 months postpartum were increased by 1.8 (ig/dL
per 1 |ig/L milk Pb at 1 month postpartum (p < 0.0001, r = 0.3). However, milk Pb only accounted for
30% of the variability in the infants" blood Pb concentrations (PbB). Using the IEUBK v2.0, a maximum
blood Pb contribution of 0.3 (ig/dL is predicted due to milk Pb at 5 months of age using a water Pb
concentration of 1.0 |ig/L (to mimic milk exposure) and an upper percentile (mean + 2SD; i.e., the 98th
percentile) intake rate of milk of 1 L/day for infants <12 months of age based on Chapter 15 of the
U.S. EPA Exposure Factors Handbook (U.S. EPA. 2011). The importance of hand Pb contamination to
BLL was investigated by Simon et al. (2007). who found BLLs of 13 infants decreased for 30-90 days
after birth before beginning to gradually increase along with hand-wipe Pb concentrations (which
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increases the likelihood of ingestion during hand-to-mouth behavior) of the infants. The infants" BLLs
were correlated with hand-wipe Pb concentrations of both the infants (r = 0.72, p < 0.01) and mothers
(r = 0.62, p < 0 .01). Thus, the contribution of breast milk Pb itself to infants" blood Pb may be
overestimated Ettinger et al. (2014) due to exposure by other pathways such as the hand-to-mouth
behavior of infants.
2.1.3.4 Exposure to Pb in Consumer Products
Consumer products have been identified as a source of Pb exposure in previous reviews (U.S.
EPA, 2013, 2006). This subsection builds upon discussions from the 2013 Pb ISA (U.S. EPA, 2013),
highlighting consumer products found in the recent literature, detailing their corresponding Pb content,
and further summarizing trends since that ISA (U.S. EPA, 2013).
Table 2-6 shows Pb content found in several consumer products, including spices, traditional
medicines, cosmetics, toys/baby products, pottery, and tobacco. This table is an update of Table 3-7 in the
2013 Pb ISA (U.S. EPA, 2013).
Although products purchased in the United States have been found with detectable Pb content,
most consumer products identified with detectable Pb content originate in countries abroad (Table 2-6). In
many cases, these products were purchased outside the United States and brought into the country by the
consumer. For example, in an analysis of 1,496 of the consumer products sampled by the New York City
Department of Health and Mental Hygiene (DOHMH) between 2008 and 2017, Hore et al. (2019)
reported 45% of spices purchased abroad had Pb content above 2 ppm, as opposed to 13% of sampled
spices purchased in the United States. The 2-ppm threshold was the permissible limit in food additives
used by DOHMH as a guidance limit (Hore et al., 2019). In 2022 the New York State Department of
Agriculture and Markets Division of Food Safety and Inspection lowered their Class II action level to
>0.21 ppm from the 1.0 ppm level set in 2016 (Ishida et al.. 2022).
Table 2-6
Pb content in various consumer products
Product
Category
Product
Location of
Purchase
Country of
Origin
Pb Content (units)
Reference
Cosmetics
Lipsticks
United States
(California)
Average:
0.36 ± 0.39 |jg/g;
Maximum: 1.32 |jg/g
Liu et al. (2013)
Costume cosmetics
United States
(California)
China,
United
States,
Taiwan
N.D. to 27 mg/kg
Perez et al. (2017)
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Product Product Location of Country of Pb Content (units) Reference
Category Purchase Origin
Lip products (lip balms,
lip glosses, lipsticks)
Online; China
(Harbin)
NR
2.48-18.22 mg/kg
Gao et al. (2018b)
Lipsticks
Iraq
NR
0.45-48.59 |jg/g
Savvadi and loannidu
(2015)
Pottery
Glazed containers
Mexico
0.026-68.6 mg/kg
Bahena et al. (2017)
Spices
Georgian saffron
United States
(New York),
Georgia
Georgia
Geo Mean:
240.1 |jg/g
Hore et al. (2019)
Tobacco
Cigarettes (filler
tobacco, filter, and ash)
Ireland
0.378-1.16 |jg/
cigarette
Afridi et al. (2015)
Cigarettes
Nigeria
Nigeria
Filler Tobacco:
17.21-74.78 |jg/g;
Filter: 4.09-
13.78 |jg/g
Benson et al. (2017a)
Cigarettes
Portugal
NR
0.44-0.72 (mean:
0.55) (pg/g)
Pinto et al. (2017)
Dried tobacco leaves
United States
Thailand
36.12 ppm (pg/g)
El Zahran et al.
(2018)
Cigarettes
China
China
Mean: 2.7718 pg/g
Li et al. (2020)
Toys and
Baby
Products
Teethers and feeding
teats
Europe
(Unspecified,
products were
made in China)
N.D. to 27.31 pg/g
Aboel Dahab et al.
(2016)
Diaper powder
United States
620,000-
639,500 pg/g
Karwowski et al.
(2017)
Children's toys and
jewelry (metallic toys
and jewelry, plastic toys,
paper/wood toys,
brittle/pliable toys, and
paint coating from toys)
China (Nanjing)
NR
0.08-860,000 mg/kg
Cui et al. (2015)
Traditional
Medicine
Ayurvedic medications:
Mahayogaraj Guggulu
(MG), Bruhat Vata
Chintamani Rasa
(BVCR)
United States
(Wisconsin);
Produced in
India and
purchased
online
MG: 48,700 mg/kg;
BVCR: 16.4 mg/kg
Meiman et al. (2015)
Kajal (eye cosmetic)
Afghanistan,
brought into
United States
Afghanistan
540,000 pg/g
CDC (2013)
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Product Product Location of Country of Pb Content (units) Reference
Category Purchase Origin
DawTway Not specified: Myanmar Median: 520 |jg/g Ritchev et al. (2011)
Myanmar or
United States
Oral Ayurvedic United States 7.3-24,000 |jg/g Hore et al. (2012)
medications (Pregnita, (New York)
Vatvidhwansan Ras,
Kankayan Bati (Gulma),
Garbhaoal Ras, Ovarin,
Garbha Dharak Yog,
Laxmana Louh, Garbha
Chintamani Ras (Vrihat)
(Swarna Yukt),
Pigmento)
Sindoor United States India U.S. Samples: Geo Shah et al. (2017)
(New Jersey); mean (SD): 5.4 (1.6)
India Max: >300,000; India
samples: 28.1 (32.4)
Max: >300,000
(pg/g)
BVCR = Bruhat Vata Chintamani Rasa; MG = Mahayogaraj Guggulu; N.D. = not detected; NR = not reported; Pb = lead;
SD = standard deviation.
Research evidence has shown exposure to cigarette smoke through use of cigarettes or
secondhand smoke can lead to increased BLLs. Richter et al. (2013) analyzed BLLs of NHANES data
from 1999 to 2008 for participants 3 years and older who responded to questions about smoking
(n = 43,627). The authors found the BLLs were higher for smokers and nonsmokers exposed to
secondhand smoke even after controlling for age of housing and occupational exposure to Pb. Apostolou
et al. (2012) analyzed BLLs and demographic information of 6,830 subjects aged 3-19 years old using
NHANES data from 1999-2004. The authors found participants in the highest quartile of serum cotinine
(>0.44 |ig/L). a biomarker indicator of recent smoke exposure, had 28% higher BLLs than those in the
lowest quartile (<0.03 (.ig/L). In addition, those living with one or two smokers had 14% and 24% higher
BLLs, respectively, than those living without smokers. Higher BLLs have also been found in studies of
Swedish smokers (Almerud et al.. 2021; Wennberg et al.. 2017). A study of the relationship between
BLLs and smoking status in an elderly population of Koreans found higher BLLs among smokers (GM:
2.09, geometric standard deviation [GSD]: 1.93 for smokers; GM: 1.90, GSD: 1.66 for nonsmokers), but
it was not statistically significant (p = 0.2597) (Lee et al.. 2017).
Electronic cigarettes are also a potential source of Pb exposure. Hess et al. (2017) measured Pb
concentrations in liquid cartridges from five brands of e-cigarettes. Pb concentrations were highly
variable, even among samples from the same brand, with median values for each brand ranging from
1,970 |ig/L to 4.98 |ig/L. Olmedo et al. (2018) and Zervas et al. (2020) also provided evidence that Pb
may be transferred from the heating coil in e-cigarettes to vapor, adding a potential source of Pb exposure
from these devices.
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2.1.3.5
Occupational Exposure
Engagement in the workplace is a common source of Pb exposure. Although its use in many
industries has decreased overtime, as of 2011, it accounted for 95% of BLLs >25 (ig/dL in adults (CDC.
2011). This section builds upon previous reviews and briefly summarizes recent studies that examine
modern occupational Pb exposure in the United States. Information on the specific Pb exposure-blood Pb
relationship in occupational cohorts can be found in Section 2.5.1.
The 2013 Pb ISA (U.S. EPA. 2013) states that operations involving Pb-containing materials in
various industries are a source of occupational Pb exposure. This occurs in many industry sectors,
including construction, manufacturing, wholesale trade, transportation, remediation, and recreation. Pilots
may be exposed to avgas through preflight fuel checks in which 75,000 to 175,000 gallons of avgas are
discarded on the ground annually. Start-up and idling of planes at high fuel-to-air ratios, along with
venting of avgas from production, transport, distribution, and storage, create additional opportunities for
exposure (NASEM. 2021). In an investigation conducted by the Wisconsin State Health Department
between 2015 and 2016, Weiss et al. (2018) found 171 (73.7%) shipyard workers had BLLs greater than
5 (ig/dL.
Ammunition is another source of occupational Pb exposure that can occur in both indoor and
outdoor firing ranges. Pb particles, dust and fumes from lead primer composed of approximately 35%
lead styphnate and lead peroxide, and bullet fragments are ejected from a gun barrel at high pressure
when a firearm is discharged. Pb exposure can occur through the inhalation of Pb in this mixture. Fine
and coarse particles, together with bullet fragments, deposit on surfaces such as hands, clothing, and
surrounding soil. Interaction with these surfaces along with the handling of Pb-containing bullets can lead
to unintended ingestion of Pb as another pathway of exposure. Pb exposure through these pathways can
also occur in a recreational setting (Beaucham et al.. 2014). Firing range employees may also be exposed
through cleaning and removal of Pb from floors, targets, and ventilation systems in the case of indoor
firing ranges (Laidlaw et al.. 2017a).
Numerous studies have shown the connection between EBLLs and ammunition. The
Occupational Safety and Health Administration (OSHA) has standards in place for firing ranges (OSHA.
2020). and the Department of Energy (DOE) has set guidelines for range design to mitigate airborne Pb
exposure (DOE. 2012). Recently, a Department of Defense (DoD)-commissioned U.S. National Academy
of Sciences report (NRC. 2013) on health effects of Pb exposure at firing ranges concluded the OSHA
standard of 40 (ig/dL is insufficient, which led to a DoD technical report on new health-based blood Pb
guidelines (U.S. APHC. 2014). The DoD released a new BLL standard in 2017 for DoD civilian
employees and service members, whether Pb exposure occurred during weapons training or during other
tasks. The standard considers BLLs at different levels of risk, where action is required for BLLs above
10 (ig/dL (U.S. APHC. 2014). A few recent studies examining the relationship between shooting ranges
and Pb exposure are highlighted below.
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Greenberg et al. (2016) took blood Pb and continuous personal airborne Pb measurements of 175
soldiers across four infantry units of the Israeli Defense Force (IDF) in a cross-sectional study. BLL
measurements were taken before and after basic and advanced training courses conducted in outdoor
shooting ranges. Soldiers (n = 174) were found to have nondetectable BLLs before basic and advanced
training courses. The percentage of soldiers with detectable BLLs increased from 21% after the basic
training course to 89% after the advanced training course. Weber et al. (2020) conducted a study to
characterize Pb exposure during five training tasks of a 45-day advanced urban assault class conducted in
both 2014 and 2016. Pb-free ammunition was used in the 2016 class, and that significantly reduced air
sampling and BLL measurements. However, mean and maximum personal air measurements were found
to be above the OSHA permissible exposure limit of 0.050 mg/m3 in both years. The authors suggest this
may be due to influences of other sources of Pb exposure, such as residual Pb present in weapons or
resuspension of Pb present in range soil due to historical usage.
In addition to personal exposure, working with Pb-containing materials or in a Pb-contaminated
work setting may also result in take-home exposures, where a worker contaminates their home
environment with Pb originating from the workplace and potentially exposes other members of the
household. Several CDC reports have found evidence of take-home Pb exposures. In 1998, an
investigation of Pb poisoning in six furniture workers and their families was performed. A father working
for a company that refinished antique furniture had a BLL of 46 (ig/dL, while the 18-year-old child and 4-
month-old daughter had BLLs of 26 (ig/dL and 24 (ig/dL, respectively. Among the families of the total of
six workers investigated, five of the six family members aged 7-12 did not have EBLLs but a 7-month-
old infant whose father's BLL was above 40 (ig/dL had a BLL of 16 (ig/dL, and it was 15 (ig/dL 30 days
later. Workers" BLLs decreased on average by 15 (ig/dL in about 3 months after an occupational Pb
safety program was put in place. In one case, a wipe sample of the carpet where a worker played with his
children was 30 (ig/ft2 but was reduced to 14 |ig/ft2 after steam cleaning; however, the 4-month-old's BLL
only decreased steadily after Pb-painted surfaces within the home were remediated (CDC. 2001).
From 2010 to 2011 it was determined that among 78 families of workers in a battery recycling
facility, 11 children (16% of 68 children <6 years of age) had confirmed BLLs >10 (ig/dL, and 39
children (57%) had BLLs >5 (ig/dL. It was also found that 85% of vehicle dust samples and 49% of house
dust samples were >40 |ig/ft2. Children's BLLs decreased 9.9 (ig/dL on average after U.S. EPA began
clean-up of employee homes and vehicles. In addition, the company was required to setup shower
facilities, shoe washes, and clean changing areas, and children with BLLs >5 (ig/dL were enrolled in case
management (CDC. 2012). In 2008, 55 new cases of EBLLs (>15 (ig/dL) in venous samples among
children <6 years were identified in Maine. No Pb-based paint or elevated Pb levels were found in the
homes of six children. It was found that these children were exposed to high Pb levels in the vehicles and
child safety seats, likely as a result of household contacts who worked in environments with high-risk to
Pb exposure (CDC. 2009).
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Ceballosa et al. (2021) examined the relationship between sociodemographic-, work-, and home-
related factors and Pb concentrations in house dust sampled from the homes of 23 construction, five
janitorial, and two autobody workers in Boston, MA. Factors from all three categories were found to be
associated with Pb in house dust, pointing to overlapping vulnerabilities. Pb in homes" dust ranged from
20 to 8,310 ppm, with those in construction workers" homes on average higher and more variable (mean:
775 ppm, median: 264 ppm, max: 8,300 ppm) than the homes of autobody and janitorial workers (mean:
296 ppm, median: 303 ppm, max: 579 ppm) suggesting that some construction workers were at risk for
take-home exposure. Other specific factors that were predictive of greater Pb concentrations in house dust
were not having a locker at work, not changing clothes after work, not washing hands after work, washing
clothes at a laundromat, and having a house built before 1978.
Other literature have also studied the link between take-home exposures and BLLs in house
occupants. Newman et al. (2015) linked a case of child Pb poisoning to a father's take-home occupational
exposure at an e-scrap recycler company. In June 2010, a male 1-year-old child and female 2-year-old
child had EBLLs of 18 (ig/dL and 14 (ig/dL, respectively, while the father's BLL was measured as
25 (ig/dL. The father did not wear personal protective equipment at work and the family reported that he
often had visible dust in his hair upon returning home. In addition, a lead risk assessment found detectable
Pb dust on the floor of the home, but no Pb paint was found. The father left his occupation after the
EBLLs were recognized, and the children's BLLs dropped to 8.7 (ig/dL (male) and 7.9 (ig/dL (female),
respectively, over the course of 3 months. Rinskv et al. (2018) characterized BLLs among employees at a
lead oxide manufacturing facility and children living in their households. Among those who worked in
the manufacturing area, average maximum BLLs consistently ranged from 40 to 59 (ig/dL. Of the 17
children examined, three had BLLs 5-9 (ig/dL and two had BLLs 10-19 (ig/dL. The researchers found
many inconsistencies in adherence to personal protective equipment use and personal hygiene protocols
which likely resulted in both occupational exposure in workers and take-home exposure in household
members. Becker et al. (2022) found no association of BLLs with paraoccupational Pb samples from dust
on the father's workpants and work shoes when using isotopic ratio analysis. Instead, other sources were
identified including paint chips, soil, house dust, turmeric, or another unknown source, depending on the
household. However, this study did not focus on occupations known to have a high risk of Pb exposure.
2.1.4 Co-Contaminants Commonly Present with Pb
Pb is commonly present in the environment with other contaminants, but the quantity and species
of these contaminants depend on the source type and environmental media in which Pb is contained. For
example, exposure to Pb associated with avgas may occur either through contact with liquid fuel and fluid
vapors during aircraft fueling operations or through inhalation of piston-engine aircraft exhaust. Exposure
to co-contaminants found in avgas is expected to differ between unused fuel and exhaust as combustion
changes the chemical properties of the fuel. Lovestead and Bruno (2009) investigated the composition of
100LL avgas and found that in addition to tetraethyl Pb, the fluid was composed of various branched and
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linear alkanes (largely isomers of hexane and pentane) and a small amount of toluene. This result was
similar to the composition disclosed by the manufacturer in the Material Safety Data Sheet. Turgut et al.
(2020) provided data on the gaseous and PMio emissions from a single reciprocating engine aircraft
fueled by 100LL avgas. A total of 70 PM samples were analyzed for concentrations of 48 trace elements
using ICP-MS. The PM samples were composed of 24 ± 12.8% trace metals, with the remaining mass
likely composed of black or organic carbon. Regardless of test condition, Pb was by far the most
abundant trace element (median 4.6 x 106 ng/m3), followed by Na, which was 40 times less abundant
(1.1 x 105 ng/m3). Other elements measured were reported in groups based on median ranges as shown in
Table 2-7. In the gaseous phase, CO2, CO, total hydrocarbons (HC), and NOx were sampled. In general,
there was an increase in NOx emissions with increasing engine speed and a decrease in CO and HC
emissions.
Emissions from some ongoing industrial activities (e.g., metal working, mining, Pb acid battery
manufacturing and recycling, glass and cement manufacturing) and resuspended PM deposited by
historical activity may contain Pb and other metals. Zotaetal. (2011) examined concentrations of Pb, Zn,
Cd, As, and Mn in dust samples taken from homes near a mining-impacted superfund site in Oklahoma.
Reported concentrations of Pb in house dust were 109 ± 138 |ig/g: co-contaminant concentrations may be
found in Table 2-7. Mixed metal particles measured from the emissions stack of a steel manufacturing
plant were found to be composed of Pb, Zn, K, and Na (Reinard et al.. 2007). Machemer (2004) also
investigated the composition of airborne particles originating from the basic oxygen furnace of iron and
steel manufacturing facilities. The <38 |im size fraction (used as a proxy for the respirable fraction) of
particles was largely composed of Fe, Al, Ca, Mg, Mn, and Si. In addition, the particles contained Pb at
concentrations ranging from 200 to 220 mg/kg and significant quantities of other potentially toxic metals
are reported in Table 2-7. The authors also note the particles had a strongly alkaline, potentially corrosive,
pH of 12.4.
Pb remobilized by wildfires may have a variety of inorganic and organic co-contaminants
(Boaggio et al.. 2022). Odigie and Flegal (2014) measured trace metal contents in ash collected from the
2012 Williams fire in Los Angeles, CA. In addition to Pb (7 to 42 (ig/g), they measured Co, Cu, Ni, and
Zn (concentrations shown in Table 2-7). Ghetu et al. (2022) investigated PAH concentrations in air during
wildfire events across the western United States between 2018 and 2020. They found 12 PAHs in air
associated with wildfires (reported in Table 2-7).
Road dust commonly consists primarily of organic carbon and crustal elements (S, Al, Fe, Ca, K,
Na, Mg); other, less abundant components commonly observed are related to brake and tire wear (Cu, Zn,
Sb, Ba, Pb, and S) and catalytic converters (Pt, Rh, Pd) (O'Shea et al.. 2021; Havs et al.. 2011; Lough et
al.. 2005). Havs et al. (2011) also noted 10 metals listed as U.S. EPA air toxics (Mn, Cr, Sb, Ni, Pb, As,
Co, Cd, Se, and Be) were generally enriched in PM0.1, and several biologically antagonistic suites of
metals (Cd, Cu, and V) were found in multiple PM size modes. Similarly, Pb along with crustal elements
such as Fe, Si, Ca, K, Mn, and Zn were observed in resuspended soil and dust, along with other
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potentially toxic or antagonistic components (Cr, As, Cu, V), which varied spatially depending on local
sources (Kundu and Stone. 2014).
Table 2-7 Co-contaminants in Pb sources
Source
Co-contaminants
Concentration
Reference
Piston-Engine
In < Sn < S < Ca < Al
[2.3-11.2] x 104 ng/m3
Turaut et al. (2020)
Aircraft Exhaust
[1.3-12.3] x 103 ng/m3
Cr < Zn < Ba < As < Sm < Se < V <
Mg
Sr < Mn < Cd < Ge < Ti < Ni < Cu
[1.2-9.4] x 102 ng/m3
La < Ag < Ga < Zr
7.6-33.8 ng/m3
Iron and Steel
Cr
1,500 mg/kg
Machemer (2004)
Manufacturing
Mn
18,000 mg/kg
Zn
5,500 mg/kg
Wildfire Ash
Co
3-11 pg/g
Odiqie and Fleqal (2014)
Cu
15-69 pg/g
Ni
6-15 pg/g
Zn
65-500 pg/g
Wildfire
dibenzo[e,l]pyrene, 6-methylchrysene,
Concentrations not
Ghetu et al. (2022)
Emissions
Reported
7,12-dimethylbenz[a]anthracene,
anthanthrene
5-methylchrysene, benzo[a]chrysene
naphtho[2,3-x]pyrene, naphtho[1,2-
b]fluoranthene
coronene, perylene,
dibenzo[a,l]pyrene
House Dust Near Zn
876 ± 627 pg/g
Zota et al. (2011)
Chat Piles
Cd
4.3 ± 6.8 pg/g
As
6.3 ± 9.9 pg/g
Mn
143 ± 98 pg/g
Ag = silver; Al = aluminum; As = arsenic; Ba = barium; Ca = calcium; Cd = cadmium; Co = cobalt; Cr = chromium; Cu = copper;
Ga = gallium; Ge = germanium; In = indium; La = lanthanum; Mg = magnesium; Mn = manganese; Ni = nickel; S = sulfur;
Se = selenium; Sm = samarium; Sn = tin; Sr = strontium; Ti = titanium; V = vanadium; Zn = zinc; Zr = zirconium.
Many studies that investigate Pb exposure by using biomarkers do not include other contaminants
in their analysis. However, Shim et al. (2017) analyzed NHANES blood and urine biomarker data for a
U.S. population six years of age or older from 2007 to 2012 to understand co-exposures of Pb with three
other metals: As, Cd, and Hg. For all metals, only measurements above the LOD were used. All possible
unique combinations of the metals were then selected wherein each metal concentration was at or above
the population median in blood, urine, or both. The weighted creatine-adjusted median values found in
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urine were 7.91 |ig/L (As), 0.2 |ig/L (Cd), 0.44 |ig/L (Hg), and 0.5 (ig/L (Pb). The weighted median
values found in blood were 0.27 |ig/L (Cd), 0.76 |ig/L (Hg), and 1.07 (ig/dL (Pb). Table 2-8 below shows
the prevalence of As, Cd, Pb, and Hg detected at or above median concentrations. The most commonly
occurring combinations were Cd/Pb (8.4%), Cd/Hg/Pb (10.6%), and As/Cd/Hg/Pb (22.1%).
Table 2-8 Specific unique combinations of As, Cd, Pb, and Hg detected at or
above the respective median concentrations in urine or blood
among the U.S. population 6 years and older, National Health and
Nutrition Examination Survey 2007-2012 data
Metal Combination3
Sample Nb
Prevalence0 - Weighted %
Prevalence0 - 95% Confidence
Interval
None
590
8.4d
7.0, 9.7
Pb
347
3.6
3.1, 4.2
As/Pb
236
2.2
1.8, 2.7
Cd/Pb
632
8.4
7.3, 9.5
Pb/Hg
294
3.7
CO
CO
As/Cd/Pb
381
4.4
3.7, 5.1
As/Hg/Pb
448
5.7
4.9, 6.5
Cd/Hg/Pb
696
10.6
9.3, 11.9
As/Cd/Hg/Pb
1,671
22.1
20.3, 23.9
As = arsenic; Cd = cadmium; Cr = chromium; Cu = copper; Hg = mercury; Pb = lead.
aAs and Hg represent total As and Hg.
bAII participants (n = 7,408) were tested for urinary and blood Cd, Pb, and Hg, as well as urinary As.
°Detected in blood and/or urine specimens at or above median concentrations.
dln 8.4% of the U.S. population 6 years and older, none of the four metals were detected at or above their respective population
medians in urine or blood.
Data sourced from Shim et al. (20171.
2.1.5 Exposure Disparities for Specific Populations
The 2013 Pb ISA (U.S. EPA. 2013) noted elevated or differential Pb exposure and biomarker
levels (such as blood Pb) have been shown to be statistically related to several population characteristics,
including age, sex, race and ethnicity, socioeconomic status (SES), proximity to Pb sources, and
residential factors. The 2013 Pb ISA (U.S. EPA. 2013) evaluated past research on these population
characteristics" relationship to Pb exposure and biomarker levels, including biological or intrinsic
(e.g., age, sex) and nonbiological or extrinsic (e.g., SES) factors. Evidence for increased exposure in this
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section primarily relies on studies that measured BLLs. BLLs and other biomarkers are further explored
in Sections 2.3 and 2.4.
2.1.5.1 Proximity to Sources of Airborne Pb Emissions
The 2006 Pb AQCD (U.S. EPA, 2006) found proximity to industrial sources likely contributes to
higher Pb exposures. The 2013 Pb ISA (U.S. EPA, 2013) stated the highest air Pb concentrations
measured using Pb-TSP monitoring were measured at monitors located near sources emitting Pb.
Additional evidence has shown EBLLs as a result of proximity to sources that emit Pb. Jones et al. (2010)
found neonates born near a Pb-contaminated hazardous waste site had significantly higher umbilical cord
BLLs (median: 2.2 (ig/dL, 95% CI: 1.5, 3.3 (ig/dL) compared with a reference group of neonates not
living near a potentially contaminated site (median: 1.1 (ig/dL, 95% CI: 0.8, 1.3 (ig/dL) but did not
analyze covariation between exposure and maternal characteristics, meaning that maternal characteristics
may have confounded results.
Benson et al. (2017b) assessed the relationship between airborne Pb sources and BLLs in children
aged 1 to 5 years. The authors used annual average ambient air Pb levels modeled by the U.S. EPA
National Air Toxics Assessment (NATA) for 2005 and industrial Pb releases obtained from the U.S. EPA
TRI. For TRI industrial releases, inverse distance squared weighted exposure, defined as the sum of
pounds of Pb released by each facility divided by the distance between each child and each industrial
facility squared, was calculated for each child to estimate Pb exposure from industrial releases. The
estimated median annual average ambient air Pb level was 1.77 ng/m3, and the median inverse distance
squared weighted exposure from Pb TRI facilities was 1,748 lb/mi2. Univariate analysis of unadjusted
data found Pb exposure from industrial releases was not significantly associated with children's BLLs,
whereas annual average ambient concentrations were significantly related (p < 0.01); odds ratios
indicated a 1.37 to 2.63% increase in BLLs for every 1 ng/m3 increase in annual average ambient air Pb
concentration. However, after adjusting for demographic covariates, including sex, race, age in months,
education level, percentage pre-1950 housing, poverty-income ratio (PIR), region, and survey cycle,
NATA estimated annual average ambient air Pb was no longer significantly related to BLLs, whereas a
significant association was found for industrial Pb releases (p = 0.001). In the adjusted model, a
10,000 lb/mi2 increase in inverse distance squared weighted exposure was associated with an estimated
1.13% (95% CI: 0.45%, 1.81%) increase in BLL. Brink et al. (2013) analyzed the relationship between air
Pb concentration and children's BLLs. BLL data were obtained from the Centers for Disease Control and
Prevention (CDC) Healthy Homes and Lead Poisoning Prevention Branch for 1,508 of the 3,220 U.S.
counties from 2000 to 2006. Modeled ambient concentrations of annual average airborne Pb at the county
level were obtained from 2005 NATA data. They found the highest 10% of estimated annual average air
Pb concentration included counties with total concentrations >0.00297 |ig/nr\ whereas the lowest 10%
was <0.000526 |ig/nr\ The proportion of tested children with BLLs >10 (ig/dL was 1.24% in the counties
with the highest 10% of estimated air Pb concentration, whereas the proportion with BLLs >10 (ig/dL was
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0.36% in counties with the lowest 10% of estimated air Pb. They also carried out a multivariate negative
binomial regression and found estimated air Pb concentration was significantly associated (p = 0.017)
with BLLs (% >10 |ig/dL) after adjusting for percentage of pre-1950 housing, rural classification, and
percentage of Black children by county. Brink et al. (2016) carried out a more detailed analysis to
examine the link between 2005 NATA-estimated annual average air Pb concentration and the BLLs of
children within 105 contiguous counties in Kansas. BLL data for children under 36 months was provided
through the Kansas Environmental Public Health Tracking Network for 2000-2005. It was found that the
mean estimated annual average Pb concentration was 0.00177 (ig/m3 in the 13 counties with at least one
resident child with BLL over 10 (ig/dL, and the mean Pb concentration was 0.00064 (ig/m3 in the counties
with no children with BLLs over 10 (ig/dL. No relationship between estimated NATA air Pb
concentration and mean BLL by census tract or county was found. However, a multilevel model to predict
BLL using distance from a Pb-emitting TRI site, adjusting for child's age in months, poverty rate, and
pre-1950 housing at the census tract level, found a significant (p < 0.001) inverse relationship between
mean BLL and distance from TRI site (i.e., higher BLL with decreasing distance).
Klemick et al. (2020) analyzed the impact of Superfund cleanup on children's BLLs by using
BLL data from the mid-1990s to mid-2010s for children aged 6 months to 5 years old residing within
5 km of at least one Superfund site in six states. They also used cleanup milestone dates from U.S. EPA's
Superfund Enterprise Management System to identify when construction was complete, which included
87 Superfund sites where Pb was identified as a contaminant. The results of their model showed that prior
to the start of cleanup, the rate of BLLs >3 (ig/dL was 4%-8% higher for children living within 2 km of
Pb-contaminated Superfund sites than those living 2-5 km away, and the difference was significant
(p<0.01).
2.1.5.2 Age
In the 2013 Pb ISA (U.S. EPA. 2013). children were concluded to have higher risk of Pb
exposure compared with adults because of hand-to-mouth contact, crawling, and poor hand washing. As
discussed in Section 2.3, children also have a higher rate of bone turnover, a higher percentage of total
body burden found in the bloodstream, and a lower overall body mass (Barry. 1975). Wang et al. (2021)
analyzed the BLLs of 68,877 participants using NHANES data from 1996 to 2016. The authors analyzed
the data both by cross-sectional analysis and birth cohort analysis. BLL data for each NHANES cycle
displayed a "U" shaped curve, with the highest BLLs among young children (1-5 years) and older adults
(>70 years), and the lowest BLLs among individuals 12-19 years old. When data were stratified into birth
cohorts, BLLs were highest in young children (age 1-5) and decreased monotonically with age. However,
the authors note that from 1999 to 2016, the 1990s cohort BLLs declined faster than other cohorts
observed; in addition, they found the rate of BLL decrease was faster before the ages of 13-17 years than
after, which may be due to the fast growth of blood volume within children, which can dilute blood Pb
concentrations.
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Jones et al. (2009) evaluated trends in 1- to 5-year-old children's BLLs based on 1988-2004
NHANES data. Their model indicated 1-to 2-year-old children were significantly more likely
(p < 0.0001) to have BLLs >10 (ig/dL than those in the 3- to 5-year age range after adjusting for
percentage of non-Hispanic Black children, having a PIR of >1.3, and living in a moderate-risk [built
~ 1950—1977] or high-risk [built before 1950] house. For the most recent NHANES cycle included in the
study (1999-2004), 2.4% (95% CI: 1.4, 3.5) of children 1-2 years old and 0.9% (95% CI: 0.4, 1.5) of
children 3-5 years old had BLLs >10 (ig/dL. This result implies there is a shift in distribution of BLLs as
young children age, likely due to differences in exposure, including behavioral influences as well as the
previously discussed age-related changes in physiology.
Senior populations may have elevated lifetime Pb exposures due to exposures that occurred prior
to the removal of leaded gasoline and broader Pb regulation. NHANES data for 2009-2010 presented in
the 2013 Pb ISA (U.S. EPA. 2013) revealed that BLLs were higher for participants 60 years or older
compared with younger adults and adolescents. Several studies found statistically significant relationships
between age and blood or bone Pb (Miranda et al.. 2010; Thcppcang et al.. 2008; Nriagu et al.. 2006). Jain
(2016) investigated the BLLs of men and women over 65 years old using NHANES data from 2003 to
2012. The authors found that those over 65 had higher BLLs than 20-64-year-olds (i.e., 28% higher
unadjusted GM, 26% higher adjusted GM). Vearrier and Greenberg (2012) analyzed 2,168 BLLs of
people over 80 from NHANES data between 1999-2010 and measured the BLLs of 76 people 80 years or
older who presented to an inner-city emergency department. GMs of NHANES-obtained BLLs ranged
from 2.66 (ig/dL (1999-2000) to 1.98 (ig/dL (2009-2010), with a decreasing trend overall. The GM of
inner-city subjects was 1.72 (ig/dL. The authors acknowledge that decreasing Pb exposure overtime may
have led to lower BLLs in the more recent data from older subjects.
2.1.5.3 Immigrant Populations
Both premigration and postmigration factors may contribute to EBLLs among U.S. immigrant
populations. Commonly identified premigration factors include immigration from areas with a high risk
of Pb exposure and use of cultural products with high Pb content (see Section 2.1.3.4). Postmigration
BLL increases are commonly caused by continued use of high-Pb cultural products as well as household
exposure to Pb in peeling paint and drinking water. BLLs in refugee children are particularly well studied
because testing is included in medical screenings conducted within 90 days of arrival to the United States
and may be followed up with a repeat exam several (typically 3-6) months later. Balza et al. (2022)
conducted a systematic review of 13 studies published between 2011 and 2021 that reported BLLs in
refugee children (<18 years of age) and compared these to either BLLs measured in the general
population or the CDC reference value. The percentage of refugee children with EBLLs reported in the 13
studies is summarized in Table 2-9. Twelve of the studies used data from entrance and/or follow-up
medical exams, whereas Ritchev et al. (2011) reported BLLs for refugees with varying time since
immigration as well as children born in the United States to refugee parents. In addition, 11 of the studies
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reported BLLs from refugees originating from many countries, whereas Ritchev et al. (2011) reported
BLLs from only Burmese refugees and Seifti et al. (2020) focused solely on Cuban refugees. Other
important variations between the studies were limits for children's age, which ranged from <6 years to
<19 years; reference levels at which BLLs were considered elevated (either 5 or 10 (ig/dL, although some
studies reported both); and sampling method (i.e., venous versus capillary). There is evidence that
capillary blood samples may be at a higher risk of contamination and biased higher compared with venous
samples. This is discussed in detail in Section 2.3.2.
Table 2-9 Prevalence of elevated blood Pb levels in refugee children
Study
Age
n
Study
Yr
EBLL
Reference
Value
Initial
EBLL
(%)
Follow-up
EBLL (%)
Other
EBLL
(%)*
Sampling Method
Ritchev et al.
(2011)c
<6
197
2009
5 |jg/dL
10 |jg/dL
37%
7.1%
Capillary; positives
confirmed with venous
draws.
Eisenberq et
al. (2011)
<7
1,148
2000-
2007
10 |jg/dL
16%
5% of initial samples and
32% of follow-up
samples were capillary.
Williams et al.
(2012)°
<6
257
2008-
2011
5 |jg/dL
10 |jg/dL
39%
9%
Not reported.
Raymond et
al. (2013)a
<16
1,007
1995-
2010
10 |jg/dL
22.7%
Both venous and
capillary used. Capillary
confirmed with second
test.
Yun et al.
(2016)
<19
8,148
2006-
2012
5 |jg/dL
21.1%
Not reported.
Sandell et al.
(2017)b
<18
225
199
2007-
2009,
2013
9 |jg/dL
5.6%
7.8%
Not reported.
Kotev et al.
(2018)
<15
1,950
2012-
2016
5 |jg/dL
11.2%
Not reported.
Geltman et al.
(2019)c
<7
3,054
1998-
2015
5 |jg/dL
10 |jg/dL
41.9%
7.9%
Venous only.
Shakva and
Bhatta (2019)c
<18
5,661
2009-
2016
5 |jg/dL
10 |jg/dL
22.3%
2.1%
Capillary; positives
confirmed with venous
draws.
Pezzi et al.
(2019)
<16
27,284
2010-
2014
5 |jg/dL
19.0%
22.7%
28% of initial samples
and 24% of follow-up
samples were capillary.
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Study
Age
n
Study
Yr
EBLL
Reference
Value
Initial
EBLL
(%)
Follow-up
EBLL (%)
Other
EBLL
(%)*
Sampling Method
Lupone et al.
(20201°
<16
705
2012-
2017
5 |jg/dL
10 |jg/dL
17%
Venous only.
Seaqle et al.
(20201°
<16
1,178
2010-
2015
5 |jg/dL
10 |jg/dL
8%
0.8%
Not reported.
Seifu et al.
(20201d
<16
301
2003-
2016
5 |jg/dL
10 |jg/dL
41.9%
7.9%
Venous only.
EBLL = elevated blood lead level; yr = year(s).
"'Other EBLL (%)" includes data from studies that did not use BLLs from entrance or follow-up medical exams.
aStudy measured BLLs in Manchester, NH (n = 639) and Providence, Rl (n = 368).
bStudy followed two cohorts of children during two different time periods (2007-2009 and 2013).
°Study reported data for both 5 |jg/dL and 10 |jg/dL reference values.
dStudy reported BLL as elevated at 10 |jg/dL for testing done prior to June 2012 and 5 |jg/dL for testing done afterward.
Only two studies, Seagle et al. (2020) and Sandell et al. (2017). reported a prevalence of EBLLs
in refugee children that was <10% of total tested children. Reported values in other studies ranged from
11.2% to 41.9%. Seven studies examined prevalence of EBLLs in refugee populations compared with a
nonrefugee comparison group. In all instances, the percentage of refugees with EBLLs was significantly
larger than the comparison group, ranging from 6 to 31 times the percentage of nonrefugee population
with EBLLs.
Several of the included studies examined the relationship between prevalence of EBLL and the
refugee's country or region of origin. Eisenberg et al. (2011) reported the prevalence of EBLLs for
children arriving in Massachusetts from Africa and West Africa were 3.8 times and 5.6 times,
respectively, higher than that of children from Europe/Central Asia (reference group). In addition,
children born in the Near East and South Asia region had a 3.6 times greater prevalence of EBLLs than
children from Europe/Central Asia. Geltman et al. (2019) reported immigrating from Africa (OR 2.49),
East Asia and the Pacific (OR 1.98), and South-Central Asia (OR 2.47) was associated with increased risk
of EBLL compared with Europe or Eurasia. Lupone et al. (2020) reported the majority of children in their
study with EBLLs arrived from countries in Africa (55.0%), and the prevalence of EBLLs was 30% for
children from the Middle East, 14.2 % for children from Southeast Asia, and 0.8% for children from
Eastern Europe. Yun et al. (2016) reported prevalence of EBLLs for children from Bhutan (26.8%),
Burma (via Thailand) (1.9%), Burma (via Malaysia) (10.5%), the Democratic Republic of the Congo
(25%), Ethiopia (13.1%), Iraq (19.9%), and Somalia (19.8%). Shakva and Bhatta (2019) reported a high
prevalence of EBLLs in children from South Asia, including Afghanistan (56.2%), Nepal (44%), Bhutan
(32.8%), and Burma (31.8%). Pezzi et al. (2019) reported a high prevalence of EBLLs for children from
India (57.9%), Afghanistan (55.1%), Burma (37.2%), Nepal (27.5%), and Syria (22.7%).
In two of the studies, time away from the country of origin correlated with lower BLLs. Kotev et
al. (2018) reported an inverse association between length of time from resettlement to testing and EBLL,
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and Shakva and Bhatta (2019) also reported a decrease in BLL with increased time since arrival.
However, time away from country of origin did not result in lowered BLLs for all children observed.
Williams et al. (2012) focused on BLLs measured during follow-up tests conducted several months post
immigration. They reported 22 children experienced a 2 (ig/dL or more increase in BLLs between two
screens. Eleven of these children had relocated to a secondary housing placement since their initial
screening test, and they attribute the increased BLLs to exposures at the new address. However, the
remaining 11 children experienced an increase in their BLLs while remaining at their initial housing
placement. Three children experienced a BLL increase above 10 (ig/dL (one from 20 to 25 (.ig/dL). The
remaining eight children experienced an increase in BLLs, but the results of the screening test remained
below 10 (ig/dL. Additional studies in the review also looked at correlations to housing and found
similarly mixed results. Eisenberg et al. (2011) found that residing in a census tract with older housing
was associated with higher BLL increases after resettlement, and Kotev et al. (2018) reported a 10-year
increase in the age of housing was associated with a 27% increase in the odds of an EBLL. However,
Ritchev et al. (2011) did not identify Pb paint or other environmental factors as significantly related to
BLLs, and Raymond et al. (2013) did not find a significant association between age of housing and BLLs
in refugees residing in either Manchester, NH or Providence, RI. However, they did find BLLs were
generally higher for refugee children than nonrefugee children living in the same buildings in Manchester
but did not find this in Providence.
Disparities in prevalence of EBLLs and increase of BLL after immigration to the United States
may be partly explained by differences in lifestyle habits and use of cultural products among refugee
populations. Ritchev et al. (2011) examined EBLLs among children of Burmese refugees in Indiana,
including U.S.-born children. They found EBLL in this population was significantly predicted by daily
use of thanakha and Daw Tway, a culturally specific cosmetic and digestive remedy, respectively.
Laboratory testing confirmed high concentrations of Pb (median 520 ppm) in Daw Tway. Differences
between nonrefugee immigrant populations have also been observed. Kaplowitz et al. (2016) investigated
a population of U.S.-born children in Michigan and compared BLLs in children with immigrant mothers
to BLLs in children with U.S.-born mothers. After controlling for individual, family, and neighborhood
characteristics, only children of South Asian-born mothers had BLLs statistically significantly higher than
children of U.S.-born mothers, whereas children of African- and Latin American-born mothers had BLLs
that did not statistically differ from children of U.S.-born mothers.
2.1.5.4 Race/Ethnicity
Both SES and race/ethnicity have been reported as correlated with BLLs. In some cases, these
factors may be linked, such as in the case of racial/ethnic minorities and those of low SES being more
likely to live in older housing (Leech et al.. 2016). Race/ethnicity may also represent a surrogate
measurement for extrinsic place-level factors that can lead to increased Pb exposure, such as urbanicity
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(Laidlaw et al.. 2023). The 2006 Pb AQCD (U.S. EPA. 2006) and the 2013 Pb ISA (U.S. EPA. 2013)
found higher blood and bone Pb levels among African Americans.
Campanella and Mielke (2008) found differences in potential exposure between racial/ethnic
groups in metropolitan New Orleans. In census blocks where surface soil Pb levels were less than
20 mg/kg, the population was 36% Black, 55% white, 3.0% Asian, and 6.0% Hispanic, based on the 2000
census. In contrast, they found that for census blocks in which soil Pb levels were between 1,000 and
5,000 mg/kg, the population was 62% Black, 34% white, 1% Asian, and 4% Hispanic. Cassidy-Bushrow
et al. (2017) measured Pb levels in tooth-matrix biomarkers among 71 children born between September
2003 and December 2007 in Detroit, MI. They found African American children had 2.2 times higher Pb
levels in the second and third trimesters (p < 0.001) and 1.9 times higher Pb levels postnatally in the first
year of life (p = 0.003) than white children. More information on using tooth-matrix biomarkers can be
found in Section 2.3.4.2.
NHANES data in Section 2.4.1 show on a national scale that non-Hispanic Black people had
higher BLLs than the average BLLs for all groups from 2011 to 2018 but were lower than non-Hispanic
whites in some years. Asian people were the racial/ethnic group with the highest BLLs from 2011 to
2018. Additional NHANES data found in Figure 2-14 show that for age groups 1-5 years and 6-10 years
GM BLLs of non-Hispanic Black children fell at a faster rate relative to other groups from 1999 to 2018.
Jones et al. (2009) compared BLLs across racial/ethnic groups using 1988-1991 and 1999-2004
NHANES data and found that although the differences between racial/ethnic groups in the percent of
BLLs >2.5 (ig/dL declined overtime, non-Hispanic Black children still had higher percentages of BLLs
>2.5 (ig/dL compared with non-Hispanic whites and Mexican Americans, with large observable
differences for BLLs between 2.5 and <10 (ig/dL. Teye et al. (2021) investigated the BLLs of 6,772
children using NHANES data from 1999 to 2016 and found that although BLLs declined for all
racial/ethnic groups over time, BLLs of non-Hispanic Black children were statistically significantly
(p < 0.05) higher than non-Hispanic white children from 1999 to 2014. However, the authors also show
that from 1999 to 2016, the gap between non-Hispanic Black children and non-Hispanic white children
decreased (a difference of 0.92 (ig/dL in 1999-2000 versus 0.15 (ig/dL in 2015-2016 data). Egan et al.
(2021) analyzed BLLs of 27,122 children from NHANES data spanning 1976-2016. The authors found
that among children 1-5 years old and 6-11 years old, non-Hispanic white children had GM BLLs lower
than Mexican American or non-Hispanic Black children for most years.
Aelion and Davis (2019) analyzed BLL data of approximately 177,000 South Carolina children
less than 6 years of age, reported between January 2011 and December 2016. Other demographic
variables including age and sex were recovered at the individual and census-block level, and the child's
residence at the block group level was also used. The mean BLL for urban block groups was found to be
statistically significantly higher than rural block groups (p < 0.0001; 2.21 versus 2.11 (ig/dL). Rural
children <1 year of age had lower BLLs than urban children of the same age (1.50 versus 2.10 (ig/dL).
Black children had statistically significantly higher mean BLLs in the urban block group (p < 0.0001;
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2.23 (ig/dL) than Black children in the rural block group (2.08 (ig/dL). Past research has supported the
idea that urban areas will have higher exposures than rural areas due to proximity to point and nonpoint
sources. However, the authors acknowledge further research is needed to identify the differential sources
that contribute to Pb exposure in early life.
Moody et al. (2016) examined disparities in children's BLLs related to race and socioeconomic
characteristics of place of residence. This study used 216,101 BLL records obtained from a statewide
database for children <1 month old to 16 years of age, taken between 2006 and 2010 in Detroit, MI.
Using bivariate regression, they determined there was an increase in mean BLLs as neighborhood SES
declined for all races. In addition, they found race was a factor regardless of SES as evidenced by higher
mean BLLs for Black children compared with white children residing in the same neighborhoods. Lynch
and Meier (2020) conducted a similar investigation, which focused on the intersectional effect of poverty,
home ownership, and racial/ethnic composition on childhood Pb exposure. This study analyzed 48,393
BLLs of children <6 years of age obtained from the Wisconsin Department of Health Services from 2014
to 2016. The samples were aggregated by 215 Milwaukee census tracts, with 225 individual childhood
blood Pb observations contributing to census tract-level means on average. They found that EBLLs were
significantly (p < 0.0001) related to predominantly low home ownership, high poverty, and majority non-
white census tracts. Further, their model showed children residing in neighborhoods with all three factors
had a 1.78 (ig/dL (95% CI: 1.44, 2.11, p < 0.0001) higher mean BLL than those in high home ownership,
low poverty, and majority white census tracts, after adjusting for average census tract housing age and
number of children. Nriagu et al. (2011) found the mean BLL among a population of 6-month- to 15-
year-old Arab American and African American children in Michigan in 2007-2008 was 3.8 ± 2.3 (ig/dL
(range: 1-18 (ig/dL) with 3.3% of the children having BLLs above 10 (ig/dL, which was higher than the
statewide average of 1.1% of children <6 years old in 2008.
2.1.5.5 SES
The 2006 Pb AQCD (U.S. EPA, 2006) found negative associations between income or other SES
metrics and blood Pb, although these relationships were not always statistically significant. Nriagu et al.
(2006) analyzed BLLs from 934 African American heads of households ranging from 14 to over
55 years of age in Detroit with household income below the 200th percentile of the federal poverty level
in 2003. They found education (p < 0.001), income (p < 0.001), and employment status (p = 0.04) were
all statistically significant predictors of BLLs, with blood Pb decreasing with some scatter as education
and income level increased. On a national level, the difference in BLLs that has historically been seen
between different income levels has been decreasing. Jones et al. (2009) evaluated the relationship
between BLLs reported for 1- to 5-year-old children in 1988-2004 NHANES data and the family's PIR,
defined as the ratio of total family income to the poverty threshold for the year of the interview. The
results of their multivariate logistic regression model found children from families with a PIR <1.3 (low
income) were significantly associated with BLLs >10 (ig/dL. However, for the most recent NHANES
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cycle included in the study (1999-2004), although the percentage of 1- to 5-year-old children having
BLLs >10 (ig/dL was higher for PIR <1.3 than for PIR >1.3 (1.8% versus 0.8%), this difference was not
statistically significant.
Wheeler et al. (2019b) investigated reasons for EBLLs (>5 (ig/dL) among children <6 years old
across 1,208 census tracts in Maryland from 2005 to 2015. They found the three statistically strongest
community predictors of EBLL in order of importance were percentage of pre-1940 housing, percentage
of African Americans in the population, and inverse median household income (meaning larger values in
disadvantaged areas) in the past 12 months. In a follow-up study on the same population, Wheeler et al.
(2022) examined the relationship between EBLLs and temporally varying neighborhood characteristics.
There were clear temporal trends in the blood Pb test data, with the percentage of EBLLs peaking in 2006
at 11% and then gradually declining to a low of 2% in 2015. As in the previous study, which did not
account for temporal variation, the percentage of pre-1940 housing and the inverse median household
income in the past 12 months were statistically important predictors of EBLL. However, this follow-up
study found the percentage of African Americans in the population was a much less important predictor
of EBLL risk than the investigators had previously observed. They also note the relationship between
these factors and risk of EBLLs remains positive and significant over the entire time period but generally
diminished over time, indicating a decline in exposure disparities. Wheeler et al. (2019a) also investigated
EBLL risk among children <6 years old in 1,332 census tracts in Minnesota from 2011 to 2015. Using a
weighted quantile sum (WQS) regression model, the five variables most significantly correlated to
children's BLLs (in order of most to least by estimated WQS index weight) were percentage of houses
built prior to 1940 (0.32), percentage not using Social Security income (0.18), percentage of housing that
was renter occupied (0.12), percentage unemployed (0.09), and percentage of African Americans in the
population (0.08). Six of the 15 variables tested were not found to be significantly predictive, including
percentage below the federal poverty level, percentage receiving public assistance income, and percentage
receiving public assistance food stamps.
2.2 Kinetics
The 2013 Pb ISA (U.S. EPA, 2013) contains previously available information on the empirical
basis for understanding Pb toxicokinetics in humans. The following Sections serve as an update to that
information. This empirically based information has been incorporated into mechanistic biokinetic models
that support predictions about the kinetics of Pb in blood and other selected tissues. The following
Sections emphasize discussion of inorganic Pb because it comprises the dominant forms of Pb to which
humans in the United States are exposed as a result of releases of Pb to the atmosphere and historic
surface deposition of atmospheric Pb. A more detailed discussion of the toxicokinetics of organic Pb can
be found in the 2006 Pb AQCD (U.S. EPA. 2006).
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2.2.1
Absorption
The major exposure routes of Pb in humans are ingestion and inhalation. Three terms that are
commonly used to aid in understanding Pb's uptake into the body are absorption, bioavailability, and
bioaccessibility. Absorption refers to the uptake of Pb ingested or inhaled into the blood from the
respiratory or gastrointestinal (GI) tract. Bioavailability is the fraction of the amount of Pb ingested or
inhaled that enters systemic circulation. If properly measured (e.g., time-integrated blood Pb), under most
conditions, Pb bioavailability is equivalent (or nearly equivalent) to Pb absorption. The time-integrated
blood Pb (i.e., the integral of blood Pb over time) provides a useful measure of bioavailability because it
reflects both recent Pb absorption as well as contributions from Pb sequestered in soft tissue and bone.
Bioaccessibility is a measure of the physiological solubility of Pb in the respiratory or GI tract. Pb must
be bioaccessible for absorption to occur. Processes that contribute to bioaccessibility include physical
transformation of Pb particles and dissolution of Pb compounds into forms that can be absorbed
(e.g., Pb2+). Bioaccessibility is typically assessed by measuring the fraction of Pb in a sample that can be
extracted into a physiological or physiological-like solution (e.g., gastric juice or solution similar to
gastric juice).
2.2.1.1 Inhalation
Systemic absorption of Pb deposited in the respiratory tract is influenced by particle size and
solubility, as well as by the pattern of regional deposition within the respiratory tract. Particle size
influences both where particles deposit in the respiratory tract and the subsequent absorption of Pb from
particles. Particles <1 |im deposited in the bronchiolar and alveolar region can be absorbed after
extracellular dissolution or can be ingested by phagocytic cells and transported from the respiratory tract.
Larger particles (>2.5 |im) that are primarily deposited in the ciliated airways (nasopharyngeal and
tracheobronchial regions) can be transferred by mucociliary transport into the esophagus and swallowed,
thus being absorbed in the GI tract. Chapter 4 of U.S. EPA (2019c) provides a detailed discussion of
factors affecting particle deposition and retention in the human respiratory tract. Section 4.2.4 of that
document specifically addresses biological factors affecting particle deposition, such as activity level and
age with an emphasis on children. The Sections below provide information on bioaccessibility of inhaled
Pb in the lung and GI tract as a function of exposure source and particle size. Empirical estimates of blood
Pb - air Pb slopes for various populations, derived from epidemiologic studies, are summarized in
Section 2.5.1.
2.2.1.1.1 Experimental Human Exposures
Inhaled Pb-laden particles depositing in the lower respiratory tract seem to be absorbed rather
similarly and totally, regardless of chemical form (Morrow et al.. 1980; Chamberlain et al.. 1978). For
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median particle diameters ranging from 0.02 |im to 0.75 |im. Figure 6.6 of Chamberlain et al. (1978)
showed the time to 50% absorption from the lung ranged from 2 to 5 hours among differing forms of Pb
(namely, Pb nitrate, Pb oxide, and four preparations of automotive exhaust). Reaching 1% remaining by
60-80 hours, the Pb nitrate (soluble, 0.75 (mi) and very fine Pb exhaust (0.02 |im) aerosols showed the
fastest clearance from the lungs. Aerosols of Pb oxide (0.75 |im) and Pb exhaust (0.5 |im. UV exposed
and unexposed) reached 2%-3% remaining by 100 hours. A carbonaceous Pb exhaust (0.5 |im) aerosol
only reached 10% remaining by 100 hours. For 0.25 |im sodium chloride droplets containing PbCk
Morrow et al. (1980) observed 50% lung Pb retention at 11.5 hours. Based on their bi-exponential decay
function, the lung is predicted to have only 1% of deposited mass remaining by 90 hours. Absorption
half-times have been estimated for radon decay progeny in adults who inhaled aerosols of Pb and bismuth
isotopes generated from decay of ~"Rn or 222Rn. The absorption half-time for Pb from the respiratory tract
to blood was estimated to be approximately 10 hours in subjects who inhaled aerosols having an activity
median particle diameter of approximately 160 nm (range 50-500 nm) (Marsh and Birchall. 1999) and
approximately 68 minutes for aerosols that have diameters of approximately 0.3-3 nm (Butterweck et al..
2002). Given the submicron particle size of the exposure, these rates are thought to represent, primarily,
absorption from the bronchiolar and alveolar regions of the respiratory tract. The results of these
experimental studies suggest Pb kinetics in the lung are at least marginally affected by Pb form and
particle size.
2.2.1.1.2 Models of Absorption Following Inhalation
The ICRP (2017) classifies the absorption of materials from particles deposited in the respiratory
tract as Type F (fast), M (moderate), and S (slow). These rates of absorption affect how much deposited
material enters the blood. For Type F Pb-laden particles, nearly all Pb moves rapidly (within an hour) into
the blood. For the Type M and S Pb-laden particles, Pb is slowly absorbed into the blood over years and
decades, respectively. For the Type M and S particle forms, much of the deposited Pb is cleared from the
respiratory tract before it can be absorbed into the blood. Most material cleared from the respiratory tract
is swallowed, and subsequent absorption (10%-20%) may occur in the GI tract. The ICRP (1995)
recommends most elements be classified as Type M because it is the least likely to excessively over- or
underestimate dissolution and absorption into the blood.
In consideration of the experimental evidence, the ICRP (2017) recommended classifying most
forms of inhaled Pb (i.e., Pb dichloride, dibromide, difluoride, hydroxide, nitrate, and oxide) as having
Type F absorption; dissociation of Pb from particles occurs at a rate of 100 day 1 (10-minute half-time).
Type M classification for Pb was recommended in Table A17 of ICRP (2002b). The ICRP (2017) has no
specific Pb forms recommended for Type M. However, a Type S classification is recommended for
mineral dusts containing Pb with dissociation of Pb from particles at a rate of 0.0001 day 1 (7,000-day
half-time). There is a specific recommendation for radon progeny with 10% dissociation at a rate of
100 day 1 (10-minute half-time) and 90% at a rate of 1.7 day 1 (10-day half-time). For particles that have
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cleared from the respiratory tract to the GI tract, ICRP (2017) recommends GI absorption fractions of 0.2
(radon and Type F) and 0.002 (Type S).
The ICRP Pb model developed by Leggett (1993) simulates age-dependent kinetics of tissue
distribution and excretion of Pb following intakes by ingestion and inhalation (see Section 2.6 for
additional discussion). The lung absorption/elimination kinetics were based largely on the Chamberlain et
al. (1978) study results for human subjects inhaling clean (i.e., not excessively carbonaceous due to a fuel
rich mixture) automotive exhaust from combustion of fuel containing 2ll3Pb-labeled tetraethyllead. The
model assumes for submicron particles that 95% of Pb deposited in the respiratory tract will be absorbed
directly into the blood with the remaining 5% transported by mucociliary clearance to the GI tract. It was
suggested that the fraction of deposited particles cleared by mucociliary clearance would be greater for
vapors and for larger particle sizes associated with occupational exposures. The elimination of Pb from
the lungs into the blood was described by Leggett (1993) as a four-compartment exponential decay
(fraction, half-time; 0.20, 1 hour; 0.35, 3 hours; 0.35, 9 hours; and 0.10, 48 hours). Thus, absorption into
the blood is initially a rapid process with 50% absorption by 4 hours and 80% absorption by 15 hours.
The remaining Pb is more slowly absorbed into the blood with an additional 15 hours to reach 90%
absorption and 6 days to reach 99% absorption of the Pb deposited in the lung. The Leggett (1993) lung
kinetics are most appropriate for airborne Pb prior to the phase-out of leaded gasoline, in part, because the
size of airborne Pb has shifted from <2.5 |im prior to the phase-out of leaded gasoline to somewhere
between 2.5 (mi and 10 (mi after the phase-out (Cho et al.. 2011).
Two physiologically based models have been developed for assessment of occupational
exposures to Pb (Sweeney. 2021; CalEPA. 2013). The California Environmental Protection Agency
(CalEPA) assumes inhaled Pb particles deposited in the alveolar region of the lung are completely
absorbed into the blood within a day. Particles deposited in the head and tracheobronchial region were
assumed to be cleared to the GI tract, where their absorption fraction was 0.30 (i.e., 30%). This 30% GI
absorption value is a 24-hour time-weighted average absorption assuming 50% absorption over 10 hours
of fasting, 19% absorption over 10 hours with liquids between meals, 12% absorption over 2 hours with
intake with solid foods, and 2 hours in which no Pb is swallowed [see Section B.3.3 on p. 82 of CalEPA
(2013)1. On the basis of its model simulations using measured aerosol size distributions of Pb in
occupational scenarios, CalEPA found that based on the patterns of deposition and subsequent absorption,
it was appropriate to assume an overall absorption fraction of 0.30 for inhaled Pb-laden particles.
Developed for the DoD, Sweeney (2021) adopted the absorption fraction of 0.30 for inhaled Pb-laden
particles used by CalEPA (2013). In review of the DoD modeling, NASEM (2020) supported an absorbed
fraction of 0.3 from inhaled Pb as a health-protective estimate. In their recent comprehensive review and
analyses (York et al.. 2023). the 0.3 absorption fraction was confirmed to represent a plausible mid-point
coefficient derived from an expanded range of theoretical particle size distributions [versus CalEPA
(2013)1 deposited in the upper and lower regions of the respiratory tract considering intake during
sedentary and outdoor activity breathing scenarios.
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Occupational studies show absorbed fractions of Pb from inhaled Pb-laden particles are varied.
Lach et al. (2015) reported Pb particle diameters were generally in the range of 0.1 to 10 |im at an indoor
firing range. Applying the approach and parameters of CalEPA (2013). NASEM (2020) estimated an AF
of only 0.231 for Pb aerosols at the indoor firing range. A study of battery workers also supports an
absorbed fraction of roughly 0.30. Assuming the particle size distribution (mass median aerodynamic
diameter, 14.1 |im: geometric standard deviation, 1.5) and particle deposition fractions in the respiratory
tract (head, 0.971; tracheobronchial, 0.026; alveolar, 0.006) for battery workers from Table B-3 of
CalEPA (2013). an absorbed fraction of 0.31 of Pb from inhaled particles is predicted as [(GI
absorption) x (head + tracheobronchial deposition fraction) + (lung absorption) x (alveolar deposition
fraction)] (i.e., 0.31 = 0.30x(0.971 + 0.026) + 1.0 x 0.006). Using personal particle samples (open-face
sampler) collected on battery workers, Dartev et al. (2014) found bioaccessibility of Pb from particles in
simulated gastric fluid was 90% (median, n = 30) and bioaccessibility in an artificial lung lining fluid
(Hatch solution) was 5.2% (median, n = 27). Using these bioaccessibility data, the absorbed fraction is
reduced to 0.27 (i.e., 0.27 = 0.30 x 0.90 x (0.971 + 0.026) + 1.0 x 0.052 x 0.006). In this case of battery
workers, assuming an absorption fraction of 0.3 appears quite reasonable.
Brown and Diamond (2023) provide a theoretical analysis of particle dissolution following
deposition in the lung and absorption of material into the blood. The authors derive dissolution rate
constants for particles depositing in the lung based on the particle's physical diameter, density, and
solubility. Pulmonary particle burden and dissolution of particles over time was modeled using the
dissolution rate in combination with rate constants for transport of deposited particles among lung regions
(e.g., mucociliary clearance from the tracheobronchial region to the head). Each log increase particle size
(e.g., from diameter of 0.1 to 1 (mi) or log decrease in dissolution rate (units of grams dissolved per cm2
particle surface area per day) was predicted to cause a log increase in the time for particle dissolution
(e.g., from 6 days to 60 days to dissolve to some fraction of the initial particle mass). A doubling of
particle density led to a doubling of the time to dissolve any given fraction of initial particle mass
remaining. The authors reported that assuming poorly soluble particle forms will enter the blood as
quickly as highly soluble forms causes an overestimation of concentrations of dissolved compounds in
blood and other extrapulmonary tissues while also underestimating their pulmonary burden. The authors
concluded that, in addition to modeling dose rates for particle deposition into the lung, physiologically
based pharmacokinetic modeling can be improved by including particle dissolution rates. However, this
model remains to be evaluated for Pb using results from human inhalation studies.
2.2.1.1.3 Urban Exposures
A few studies have quantified the bioaccessibility in the GI tract of Pb in atmospheric particles,
based on various in vitro extraction methods. In a study of PMio and PM2.5 samples collected in February-
March 2006 from downtown Vienna, Austria, Falta et al. (2008) used synthetic gastric juice to investigate
the bioaccessibility of metals, including Pb. The Pb concentrations associated with the PM10 and PM2.5
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samples were almost identical, indicating most of the Pb was associated with fine particles. The
percentage extracted by synthetic gastric juice was, on average, 86% and 83% Pb for PM2.5 and PM10
fractions, respectively. In a similar study, Gao et al. (2018a) collected PM10 and PM2.5 samples in Harbin,
China in the summer (July, August 2014) and winter (October, November 2014). These authors evaluated
total GI bioaccessibility of Pb from particles in simulated salivary, gastric, and intestinal fluids. For the
winter samples, ambient air Pb concentrations between PM10 and PM2.5 samples were quite similar, again
indicating most of the Pb was associated with fine particles. The total percent in fluid extracts
(i.e., bioaccessible) was, on average, 22% and 23% Pb for PM2.5 and PM10 fractions, respectively.
However, for the summer samples, ambient air Pb concentrations were 2.8 times greater in PM10 than
PM2.5 samples, and the extracted fractions were quite low (4% and 1% Pb for PM2 5 and PM10 fractions,
respectively).
Several studies have examined the bioaccessibility of airborne Pb-laden particles in the lung. Two
common extraction fluids are used to simulate bioaccessibility. Gamble's solution is used to mimic the
neutral conditions (pH 7.4) of epithelial/interstitial fluids. Artificial lysosomal fluid (ALF) is used to
mimic cellular conditions that exist following an immune response in the lung, associated macrophage
activity, and acidic (pH 4.5) conditions. Wiseman and Zereini (2014) collected PM10, PM2.5 and PMi
samples between June 2009 (summer) and November 2010 (autumn) in Frankfurt, Germany at an area
affected by four-lane traffic. For PM10, PM2.5, and PMi, the average bioaccessible fraction of Pb in ALF
was 0.96, 0.84, and 0.78, respectively; in Gamble's solution, it was 0.26, 0.04, and 0.05, respectively, da
Silva et al. (2015) assessed the pulmonary bioaccessibility in simulated lung fluid (Gamble's solution) of
PM10 samples collected at four sampling stations during winter (June-July 2010) in Rio de Janeiro,
Brazil. In general, all sampling stations were in mixed residential/commercial areas with intense
automotive traffic. Similar to the fraction of 0.26 for PM10 reported by Wiseman and Zereini (2014). the
overall mean lung bioaccessible fraction of PM10 in Gamble's solution was 0.22 (range: 0.11-37). Niu et
al. (2010) determined the bioaccessibility of Pb in fine (0.1-1.0 |im) and ultrafine (<0.1 |im) urban
airborne PM from two sites in 1992-1993 and 1999-2000 within Ottawa, Canada. The bioaccessibility
was based on Pb extraction in ammonium acetate (pH = 7 to simulate the neutral lung environment). The
nano fraction accounted for 33% of Pb mass, the fine fraction was 42% of Pb mass, and the remaining
25% was associated with particles >1 (mi in size. Although the Pb concentration declined by 7% for fine
and 13% for nano particles between the initial and later sampling periods, increases in bioaccessibility
increased potentially absorbed Pb by 1.2 times (fine) and 1.5 times (nano). For the 1999-2000 sampling
phase, the bioaccessibility fraction showed clear increase with decreasing particle size (fraction, size;
0.15, 1 (mi; 0.20, 0.2 (jm; 0.28, 0.06 (mi).
Considering possible resuspension and human inhalation, Dean et al. (2017) measured the
bioaccessibility of urban street dusts collected from five northern U.K. cities: Durham, Edinburgh,
Liverpool, Newcastle upon Tyne, and Sunderland. Twenty-one samples were collected, dried,
disaggregated, and sieved to <125 |im and then < 10 |im (particle sizes likely to be inhaled by
pedestrians). Bioaccessibility was assessed in a synthetic epithelial lung fluid having a pH of 7.4. Dusts
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were added to the synthetic fluid, shaken, and maintained at 37°C for 96 hours. The authors found a
bioaccessible fraction of 4.2 ± 2.2% (range 1.2%—8.8%). These low bioaccessible fractions are consistent
with low bioaccessible fractions of 0.26, 0.04, and 0.05 reported for PMio, PM2.5, and PMi, respectively,
in Gamble's solution by Wiseman and Zereini (2014).
2.2.1.1.4 Smelting and Mining Exposures
Goix et al. (2016) reported the Pb bioaccessibility in ALF and Gamble's solution of PM0.5
samples collected from areas of smelting and mining in Oruro, Bolivia. PM0.5 samples represented 79%
and 71% of PM2.5 mass in the smelting and mining area samples, respectively. The bioaccessibility
fractions of PM0.5 from smelting samples were 0.70 (ALF) and 0.32 (Gamble's solution). Considerably
lower bioaccessibility fractions were reported for mining at only 0.07 (ALF) and 0.02 (Gamble's
solution). Gastric bioaccessibility of dust samples was also greater in the smelting than mining areas. Li et
al. (2016) assessed changes in Pb bioaccessibility in ALF and Gamble's solution of PM2.5 samples
collected before (June-July), during (August), and after (September-October) the 2014 Youth Olympic
Games in Nanjing, China. Two important Pb sources in PM2 5 from urban sites of Nanjing are coal
emissions and smelting activities, the latter of which were shut down during the Olympics (but the former
continued for electricity production). Pb bioaccessibility in ALF was lower (61 ± 4.3%, n = 9) during the
games than before (66 ± 6.4%, n = 10) or after (78 ± 4.6%, n = 13). The lower bioaccessibility in ALF
during the games may reflect the importance of the normally occurring smelting operations. Interestingly,
the average Pb bioaccessibility based on Gamble's solution was higher (20%) during games than before
(10%) or after (11%).
Xing et al. (2020) examined both ALF and gastric bioaccessibility of dust samples collected from
the exterior windowsill (i.e., the trough) of the 1st through 9th floors of buildings in Jiyuan City, Henan
Province, northern China, an area affected by Pb smelting and other industries. Trough dusts are generally
assumed reflective of outside air and exterior contamination that runs down windows with rain to collect
in the trough area. Dusts were size fractioned into <10, 10-45, and 45-125 |im. which are reflective of the
size distribution of the dust and not the airborne particle sizes that were transported and deposited on the
troughs or higher building surfaces. On the basis of isotopic ratios, the authors concluded higher floors
were more affected by smelting and lower floors more by resuspension of soils. At the four sample sites
most affected by smelters, the bioaccessible fractions in ALF (0.835) and gastric fluid (0.812) were nearly
identical. This suggests bioaccessibility in ALF may be a reasonable substitute for gastric bioaccessibility
for smelting dusts.
2.2.1.1.5 Organic Pb Exposures
Alkyl Pb compounds can exist in ambient air as vapors. Inhaled tetraalkyl Pb vapor is nearly
completely absorbed following deposition in the respiratory tract. As reported in Section 4.2.1 of the 2006
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Pb AQCD (U.S. EPA, 2006), a single exposure to vapors of radioactive (2ll3Pb) tetraethyl Pb resulted in
37% initially deposited in the respiratory tract, of which -20% was exhaled in the subsequent 48 hours
(Heard etal., 1979). In a similar experiment conducted with 203 Pb tetramethyl Pb, 51% of the inhaled
2ll3Pb dose was initially deposited in the respiratory tract, of which -40% was exhaled in 48 hours (Heard
etal.. 1979).
Estimation of bioavailability of tetraethyl Pb following combustion is relevant to some aviation
exposures (e.g., persons exposed to leaded gasoline used in piston-engine aircraft). Chamberlain et al.
(1975) suggested 35% of inhaled combustion products of tetraethyl 2ll3Pb fuel [likely to have been a
mixture dominated by inorganic Pb halides but may also have included alkyl Pb species (U.S. EPA,
2006)1 are deposited and then retained in adult lungs with a half-life of 6 hours. Fifty percent of that
2ll3Pb was detectable in the blood within 50 hours of inhalation, and the rest was found deposited in bone
or tissue. Chamberlain et al. (1975) estimated a 1 (ig/dL increment in blood Pb could result from
continuous inhalation over a period of months of a Pb-laden aerosol at a concentration of 1 |ig Pb/m3
generated by vehicle engine combustion of fuel containing tetraethyllead.
2.2.1.2 Ingestion
The extent and rate of GI absorption of ingested inorganic Pb are influenced by physiological
states of the exposed individual (e.g., age, fasting, nutritional calcium (Ca2+) and iron (Fe) status,
pregnancy) and physicochemical characteristics of the Pb-bearing material ingested (e.g., particle size,
mineralogy, solubility). Pb absorption in humans may be a capacity-limited process, in which case the
percentage of ingested Pb that is absorbed may decrease with increasing rate of Pb intake. Numerous
observations of nonlinear relationships between blood Pb concentration and Pb intake in humans provide
support for the likely existence of a saturable absorption mechanism or some other capacity-limited
process in the distribution of Pb in humans (Sherlock and Ouinn. 1986; Sherlock et al.. 1984; Pocock et
al.. 1983; Sherlock et al.. 1982). While evidence for capacity-limited processes at the level of the
intestinal epithelium is compelling, the dose at which absorption becomes appreciably limited in humans
is not known.
2.2.1.2.1 Physiologic Factors
In adults, estimates of absorption of ingested water-soluble Pb compounds (e.g., Pb chloride, Pb
nitrate, Pb acetate) range from 3 to 10% in fed subjects (Maddaloni et al.. 1998; Watson et al.. 1986;
James et al.. 1985; Heard and Chamberlain. 1982; Rabinowitz et al.. 1980). The absence of food in the GI
tract increases absorption of water-soluble Pb in adults. Reported estimates of soluble Pb absorption range
from 26% to 70% in fasted adults (Maddaloni et al.. 1998; James et al.. 1985; Blake et al.. 1983; Heard
and Chamberlain. 1982; Rabinowitz et al.. 1980). Reported fed:fasted ratios for soluble Pb absorption in
2-49
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adults range from 0.04 to 0.2 (James et al.. 1985; Blake et al.. 1983; Heard and Chamberlain. 1982;
Rabinowitz et al.. 1980).
Limited evidence demonstrates GI absorption of water-soluble Pb is higher in children than in
adults. Estimates derived from dietary balance studies conducted in infants and children (ages 2 weeks to
8 years) indicate ~40%-50% of ingested Pb is absorbed (Ziegler et al.. 1978; Alexander et al.. 1974).
Experimental studies provide further evidence for greater absorption of Pb from the gut in young animals
compared with adult animals (Aungst et al.. 1981; Kostial et al.. 1978; Pounds et al.. 1978; Forbes and
Reina. 1972). The mechanisms for an apparent age difference in GI absorption of Pb have not been
completely elucidated and may include both physiological and dietary factors (Mushak. 1991). To further
investigate the effects of the presence of food in the GI tract on Pb absorption, children (3-5 years old)
who ate breakfast had lower BLLs compared with children who did not eat breakfast (Liu et al.. 2011).
This difference persisted after controlling for nutritional variables (blood iron [Fe], calcium [Ca2+], copper
[Cu], magnesium [Mg], and zinc [Zn]). This observation may be explained by lower GI absorption of Pb
ingested with or in close temporal proximity to meals. Direct evidence for meals lowering GI absorption
of Pb has also been reported for adults (Maddaloni et al.. 1998; James et al.. 1985).
Nutritional interactions of Pb with dietary elements (e.g., Fe, Ca2+, Zn) are complex. Pb competes
with other elements for transport and binding sites that can result in adjustments of homeostatic regulators
to absorb and retain needed elements. Additionally, low levels of macronutrients may alter Pb
bioaccessibility in the GI tract. Genetic variation in absorption and metabolism may modify all of the
above.
Children who are iron deficient have higher blood Pb concentrations than similarly exposed iron-
replete children, suggesting iron deficiency may result in higher Pb absorption or, possibly, other changes
in Pb biokinetics that contribute to altered blood Pb concentrations (Schell et al.. 2004; Marcus and
Schwartz. 1987; Mahaffev and Annest. 1986). Studies conducted in animal models have provided direct
evidence for interactions between iron deficiency and increased Pb absorption, perhaps by enhancing
binding of Pb to iron-binding proteins in the intestine (Bannon et al.. 2003; Morrison and Quarterman.
1987; Barton et al.. 1978b). An analysis of data from a sample of 448 women (ages 20 to 55 years) did
not find a significant association between iron body stores (indicated from serum ferritin concentration)
and blood Pb concentrations, although depleted irons stores (serum ferritin of <12 |ig/L) were associated
with higher blood concentrations of cadmium (Cd), cobalt (Co), and manganese (Mn) (Meltzer et al..
2010). Healthy infants (97 males, 113 females; median age: 11.4 months; range: 8-23 months) underwent
iron-deficiency screenings from July 2014 to June 2016 in Seoul, South Korea (Choi et al.. 2017). The
infants had no intake of herbal medicine, iron, or zinc supplements in the prior 3 months. Iron deficiency
was associated (p < 0.001) with an increased median blood Pb concentration of 1.24 (ig/dL (interquartile
range: 0.84, 1.64) relative to no deficiency, wherein blood Pb concentration was 0.75 (ig/dL (interquartile
range: 0.51, 1.10). The presence of iron-deficiency anemia was associated (p < 0.001) with a further
increase in median blood Pb to 1.44 (ig/dL (interquartile range: 1.14, 1.80) relative to its absence, wherein
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blood Pb was 0.79 (ig/dL (interquartile range: 0.51, 1.14). In a Norwegian study (Mcltzcr et al.. 2016) of
smoking women (n = 267; mean age = 38.3 years, range: 21-55 years), no correlation was observed
between blood Pb and blood iron concentrations with either original data values (r = -0.01) or log-
transformed data (r = 0.00). The effects of iron nutritional status on blood Pb include changes in blood Pb
concentrations in association with genetic variation in genes involved in iron metabolism. For example,
genetic variants in the hemochromatosis gene (HFE) and transferrin genes are associated with higher
blood Pb concentrations in children (Hopkins et al.. 2008). In contrast, HFE gene variants are associated
with lower bone and BLLs in elderly men (Wright et al.. 2004).
Several studies have suggested dietary Ca2+ may have a protective role against Pb by decreasing
absorption of Pb in the GI tract and by decreasing the mobilization of Pb from bone stores to blood. In
experimental studies of adults, absorption of a single dose of Pb (100,300 |ig Pb chloride) was lower
when the Pb was ingested together with Ca2+ carbonate (0.2 g Ca2+ carbonate) than when the Pb was
ingested without additional Ca2+ (Blake and Mann. 1983; Heard and Chamberlain. 1982). A similar effect
of Ca2+ occurs in rats (Barton et al.. 1978a). Similarly, an inverse relationship was observed between
dietary Ca2+ intake and blood Pb concentration in children, suggesting children who are Ca2+ deficient
may absorb more Pb than Ca2+-replete children (Elias et al.. 2007; Schell et al.. 2004; Mahaffev et al..
1986; Ziegler et al.. 1978). These observations suggest Ca2+ and Pb share and may compete for common
binding and transport mechanisms in the small intestine, which are regulated in response to dietary Ca2+
and Ca2+ body stores (Fullmer and Rosen. 1990; Bronner et al.. 1986). However, animal studies have also
shown multiple aspects of Pb toxicokinetics are affected by Ca2+ nutritional status. For example, feeding
rats a Ca2+-deficient diet is associated with increased Pb absorption, decreased whole-body Pb clearance,
and increased volume of distribution of Pb (Aungst and Fung. 1985). These studies suggest associations
between Ca2+ nutrition and blood Pb that have been observed in human populations may not be solely
attributable to effects of Ca2+ nutrition on Pb absorption. Other potential mechanisms by which Ca2+
nutrition may affect blood Pb and Pb biokinetics include effects on bone mineral metabolism and renal
function.
Blood Pb concentrations in young children have also been shown to increase in association with
lower dietary Zn levels (Schell et al.. 2004). Mechanisms for how Zn affects blood Pb concentration,
(i.e., whether it involves changes in absorption or changes in distribution and/or elimination of Pb) have
not been determined.
2.2.1.2.2 Mineralogical Factors
Dissolution of Pb from the soil/mineralogical matrix in the stomach appears to be the major
process that renders soil Pb bioaccessible for absorption in the GI tract. Absorption of Pb in soils and dust
has been most extensively studied in the in vivo swine model. Gastric function of swine is thought to be
sufficiently similar to that of humans to justify use of swine as a model for assessing factors that may
affect GI absorption of Pb from soils in humans (U.S. EPA. 2021b; Juhasz et al.. 2009; U.S. EPA. 2007a;
2-51
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Casteel et al.. 2006; Casteel et al.. 1997). Other practical advantages of the swine model over rodent
models have been described and include absence of coprophagia; ease with which Pb dosing can be
administered and controlled; and higher absorption fraction of soluble Pb (e.g., Pb acetate) in swine,
which is more similar to humans than rats (Smith et al.. 2009). The swine studies measure blood and/or
tissue Pb (e.g., kidney, liver, bone) concentrations following oral dosing of swine with either Pb-laden
soil or with a highly water-soluble and fully bioaccessible form of Pb (e.g., Pb acetate). A comparison of
the internal concentrations of Pb under these two conditions provides a measure of the bioavailability
(i.e., absorption) of Pb in soil relative to that of Pb acetate, which is typically referred to as relative
bioavailability (RBA). RBA measured in the swine assay is equivalent to the ratio of the absorbed
fraction (AF) of ingested dose of soil Pb to that of water-soluble Pb acetate
(e.g., RBA = AFsoiiPb/AFpb acetate) •
U.S. EPA (2021b) provides a review of published studies conducted in swine to assess Pb RBA
in 41 different soil or "soil-like" test materials. Table 2-10 summarizes RBA data for varied forms and
sources of Pb. The mean of RBA estimates from 31 soils was 0.54 (±0.32[SD]), the median was 0.60, and
the 5th to 95th percentile range was 0.11 to 0.97. RBA estimates for soils collected from eight firing
ranges were approximately 1.0 (Bannon et al.. 2009). The relatively high RBA for the firing range soils
may reflect the high abundance of relatively unencapsulated Pb carbonate (30%-90% abundance) and Pb
oxide (160%) in these soils. Similarly, a soil sample (low Pb concentration) mixed with a National
Institute of Standards and Technology paint standard (55% Pb carbonate, 44% Pb oxide) also had a
relatively high bioavailability (0.72) (Casteel et al.. 2006). A somewhat lower RBA has been reported for
a variety of paints having an average RBA of 0.61 ± 0.24 with a large RBA range from 0.35 to 1.1 (Hunt.
2016). Samples of smelter slag, or soils in which the dominant source of Pb was smelter slag, had
relatively low RBA (0.14-0.53, three sites), as did a sample from a mine tailings pile (RBA = 0.06-0.40,
two sites) and a sample of finely ground galena mixed with soil (RBA = 0.01) (U.S. EPA. 2021b). U.S.
EPA (2021b) recommended a central tendency RBA of 0.6 (60%) for Pb in soils that are not associated
with firing ranges. This is consistent with a separate meta-analysis of soil Pb data (Dong et al.. 2016).
Table 2-10 Relative bioavailability for varied Pb forms and sources
Relative Bioavailability Pb Form
i /^-ino/ dda\ Anglesite (Pb sulfate), Fe/Pb sulfate, Galena (Pb sulfide), Pb-
LOW (< 1U /o KdA) ¦ , . ,r i.
v ' related sulfosalts
.. r,r,A> Fe/Pb oxide, Fe/Pb silicate, Mine tailings, Dust and soil (mining
Medium (20%-55% RBA) associated), PbO, Pb phosphate, Slag Zn/Pb silicate
Elevated (55 < RBA < 90%) Pb-based paint, Dust and soil (smelter associated), Urban soil
v ' (legacy leaded gasoline and atmospheric deposition)
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Relative Bioavailability
Pb Form
High (>90% RBA)
Pb ammunition (Pb shot), Cerussite (Pb carbonate), MnPb
Oxide
RBA = relative bioavailability.
Source: (Wang et al.. 2022: U.S. EPA. 2021b: Dong et al.. 2016: Goix et al.. 2016: Bannon et al.. 2009: Casteel et al.. 20061.
Drexler and Brattin (2007) developed an in vitro bioaccessibility (IVBA) assay for soil Pb that
uses extraction fluid composed of glycine, deionized water, and hydrochloric acid at a pH of 1.50 that is
combined with sieved test material (<250 |im) for 1 hour. The assay was tested for predicting in vivo
RBA of 18 soil-like test materials that were assayed in a juvenile swine assay (Casteel et al.. 2006). A
regression model relating IVBA and RBA was derived based on these data (Equation 2-1):
RBA = (0.878 X IVBA) - 0.028 Equation 2-1
where RBA and IVBA are expressed as fractions (i.e., not as percent). The weighted r2 for the
relationship (weighted for error in the IVBA and RBA estimates) was 0.924 (p < 0.001). The IVBA assay
reported in Drexler and Brattin (2007) has been identified by U.S. EPA as a validated method for
predicting RBA of Pb in soils for use in risk assessment (U.S. EPA. 2015. 2007b). A review of soil Pb
RBA estimates made using the IVBA assay described above and Equation 2-1 identified 270 estimates of
Pb RBA in soils obtained from 11 hazardous waste sites. The mean for the sitewide RBA estimates
(n = 11 sites) was 0.57 (SD 0.15), the median was 0.63, and the 5th to 95th percentile range was 0.34 to
0.71. The use of the IVBA assay for predicting in vivo RBA for soils that have been treated with
amending agents that alter the solubility or mobility of Pb, such as those that have been treated with high
levels of phosphate (e.g., 1% phosphoric acid w/w), is not recommended (U.S. EPA. 2015).
Equation 2-1 cannot be reliably extrapolated to other in vitro assays that have been developed for
estimating Pb bioaccessibility without validation against in vivo RBA measurements made on the same
test materials. Comparisons of outcomes among different in vitro assays applied to the same soil test
materials have found considerable variability in IVBA estimates (Juhasz et al.. 2011; Smith et al.. 2011;
Saikat et al.. 2007; Van de Wiele et al.. 2007). This variability has been attributed to differences in assay
conditions, including pH, liquid:soil ratios, inclusion or absence of food material, and differences in
methods used to separate dissolved and particle-bound Pb (e.g., centrifugation versus filtration). Smith et
al. (2011) found that algorithms for predicting RBA based on two different IVBA assays did not yield
similar predictions of RBA when applied to the same material. Given the dependence of IVBA outcomes
on assay conditions, in vitro assays used to predict in vivo RBA should be evaluated against in vivo RBA
estimates to quantitatively assess uncertainty in RBA predictions (U.S. EPA. 2007b).
Absorption of Pb in house dust has not been rigorously evaluated quantitatively in humans or in
experimental animal models. The RBA for paint Pb mixed with soil was reported to be approximately
0.72 (95% CI: 0.44, 0.98) in juvenile swine, suggesting paint Pb dust reaching the GI tract may be highly
bioavailable (Casteel et al.. 2006). The same material yielded a bioaccessibility value (based on IVBA
2-53
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assay) of 0.75 (Drcxlcr and Brattin. 2007). which corresponds to a predicted RBA of 0.63, based on
Equation 2-1. A review of indoor Pb RBA estimates made using the IVBA assay and Equation 2-1
identified 100 estimates of Pb RBA in dusts obtained from two hazardous waste sites. Mean Pb RBAs for
the Herculaneum site were 0.47 (SD 0.07, 10 samples) for indoor dust and 0.69 (SD 0.03, 12 samples) for
soil. At the Omaha site, mean Pb RBAs were 0.73 (SD 0.10, 90 samples) for indoor dust and 0.70 (SD
0.10, 45 samples) for soil. Yu et al. (2006) applied an IVBA method to estimate bioaccessibility of Pb in
house dust samples collected from 15 urban homes. Homes were selected for inclusion in this study based
on reporting to the state department of health of at least one child with a blood Pb concentration
>15 (ig/dL, and Pb paint dust may have contributed to indoor dust Pb. The mean IVBA was 0.65 (SD
0.08, age: 52.5 to 77.2 months).
The above results, and the IVBA assays used in studies of interior dust, have not been evaluated
against in vivo RBA estimates for dust samples. Although expectations are that a validated IVBA
methodology for soil would perform well for predicting RBA of interior dust, this validation has not
actually been experimentally confirmed. Factors that may affect in vivo predictions of RBA of interior
dust Pb could include particle size distribution of interior dust Pb and the composition of the dust matrix,
which may be quite different from that of soil.
2.2.1.2.3 Particle Size
Several studies have shown Pb concentrations in soil, bioavailability, and particle adherence to
the hands (which affects the probability of incidental ingestion) all depend on particle size. In past
reviews, studies showed GI absorption of Pb from larger Pb-containing particles (>100 |im) tended to be
lower than from smaller particles (Healv et al.. 1992; Barltrop and Meek. 1979). Stalcup (2016) reviewed
literature (January 2000-December 2011) on the relationship between particle size and dermal adherence
and between particle size and Pb enrichment. Particle size distribution of metals in shooting ranges,
incinerators, mine tailings and associated background soil samples from three mining sites, as well as
urban soils and dusts, demonstrated consistent enrichment in particle size fractions smaller than <150 |im
(Juhasz et al.. 2011; Kim et al.. 2011; Luo et al.. 2011; Madrid et al.. 2008; Ljung et al.. 2007; Pve et al..
2007; Liung et al.. 2006; Momani. 2006; Weiss et al.. 2006; Tawinteung et al.. 2005). The importance of
particle size as it relates to dermal adherence, consequent ingestion, and variance in contaminant levels
may also apply to other metals, PAHs, or other contaminants in soil and dust (Beamer et al.. 2012; Ruby
and Lownev. 2012; Bergstrom et al.. 2011; Siciliano et al.. 2009; Yamamoto et al.. 2006). More recent
studies continue to support Pb enrichment in smaller particle sizes and provide bioaccessible data as a
function of particle size.
Logiewa et al. (2020) examined metal content of road dust samples collected in three industrial
and mining towns in southern Poland. The concentration of Pb generally increased with decreasing
particle size and was greatest in particles <2 |im. Figure 2-2 illustrates the cumulative distribution of Pb
mass in the dust samples. Of note, the figure shows 73% of Pb mass is associated with particles <150 |im
2-54
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and 80% with particles <250 |im. This suggests the recently developed sieving recommendations
(Stalcup. 2016) will not negatively affect the mass of soil sampling required or the validity-established
IVBA methodology. Kama et al. (2017) specifically examined the effect of sieving <150 (mi versus
<250 |im on the determination of IVBA. They examined bioaccessibility of Pb in soils dried and sieved
into several size fractions (<250 to >150, <150 to >75, <75 to >38, and <38 |im) using a validated IVBA
technique (Stalcup. 2016; U.S. EPA. 2007b). Of the four soil types examined, only one showed an
increasing trend (r = 0.012) in IVBA Pb with decreasing soil-size fraction. The authors concluded that
sieving to <150 |im rather than <250 (mi would not undermine currently validated IVBA protocols in
future bioavailability studies.
Goix et al. (2016) reported gastric bioaccessibility of dust samples collected from areas of
smelting and mining in Oruro, Bolivia using an in vitro method validated against the in vivo juvenile
swine technique. The bioaccessible fraction was greater in dusts associated with smelting (0.63) than
mining (0.13), but no clear effect of particle size on bioaccessibility was observed. In a study of 16 soil
samples contaminated by Pb from varied sources (e.g., shooting range, incinerator, smelting/mining),
Juhasz et al. (2011) also examined the effect of particle size on bioaccessibility. In six of the 16 samples,
bioaccessibility increased with progression to finer particle sizes (<50 versus <100 versus <250 (mi) with
the largest changes being about a 25% increase (e.g., from 38 to 63%) going from <250 to <50 |im
particles. However, the largest bioaccessibility change was in the opposite direction from soil collected at
a shooting range, where bioaccessibility increased from about 69% to 99%, going from <50 to <250 (mi
particles. Overall, among studies, there are no consistent changes in IVBA as a function of particle size.
2-55
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1 1 1 1 1 1 1 1 1 1 ~
5 10 20 30 40 50 60 70 80 90 95
Cumulative Percent Pb Mass Less Than Indicated Size
Source: Points are data derived from Tables 4 and 5 of Loaiewa et al. (2020). The solid line is a log normal fitted to the data with a
median of 33 |jm and a geometric standard deviation of 11.5. Not illustrated in the figure is the 8% of Pb mass that was found
associated with particles between 1 and 2 mm.
Figure 2-2 Distribution of Pb in road dust samples collected in three
industrial and mining towns located in southern Poland.
2.2.2 Distribution and Metabolism
A simple conceptual representation of Pb distribution is that it contains a fast turnover pool,
comprising mainly soft tissue, and a slow pool, comprising mainly skeletal tissues (Rabinow itz et al..
1976). The highest soft tissue concentrations in adults occur in liver and kidney cortex (Gerhardsson et
al.. 1995; Oldereid et al.. 1993; Gerhardsson et al.. 1986; Barry. 1975; Gross et al.. 1975). Pb in blood
(i.e., plasma) exchanges with both of these compartments.
2.2.2.1 Blood
Blood comprises -1% of total Pb body burden. Pb in blood is found primarily (>99%) in the red
blood cells (RBCs) (Smith et al.. 2002; Manton et al.. 2001; Bergdahl et al.. 1999; Bergdahl et al.. 1998;
Hernandez-Avila et al.. 1998; Bergdahl et al.. 1997a; Schiitz et al.. 1996). Delta-aminolevulinic acid
dehydratase (ALAD) is the primary binding ligand for Pb in erythrocytes (Bergdahl et al.. 1998; Xie et
al.. 1998; Bergdahl et al.. 1997a; Sakai et al.. 1982). Two other Pb-binding proteins have been identified
in the RBC, a 45 kDa protein (Kmax 700 (ig/dL; Kd 5.5 |ig/L) and a smaller protein band having a
2-56
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molecular weight of <10 kDa (Bergdahl et al.. 1998; Bergdahl et al.. 1997a; Bergdahl et al.. 1996). Of the
three principal Pb-binding proteins identified in RBCs, ALAD has the strongest affinity for Pb (Bergdahl
et al.. 1998) and appears to dominate the ligand distribution of Pb (35% to 84% of total erythrocyte Pb) at
BLLs below 40 (ig/dL (Bergdahl et al.. 1998; Bergdahl et al.. 1996; Sakai et al.. 1982). Pb binding to
ALAD is saturable; the binding capacity was estimated to be -850 (ig/dL RBCs (or -40 (ig/dL whole
blood), and the apparent dissociation constant has been estimated at -1.5 |ig/L (Bergdahl et al.. 1998).
Hematocrit is somewhat higher in the neonate at birth (51%) than in later infancy (35% at 6 months),
which may lead to a decrease in the total binding capacity of blood over the first 6 months of life that
results in a redistribution of Pb among other tissues (Simon et al.. 2007).
The primary binding ligand for Pb in RBCs is encoded by a single gene that is polymorphic in
two alleles (ALAD1 and ALAD2). These can be co-dominantly expressed. Thus, three different
genotypes are possible (ALAD 1-1, ALAD 1-2, and ALAD 2-2). In the 2013 Pb ISA (U.S. EPA. 2013).
many studies showed individuals with the ALAD-2 gene had higher BLLs. However, there was also
evidence showing there was no difference in BLLs between ALAD-1 or ALAD-2 carriers or even lower
BLLs for ALAD-1-2/2-2 carriers. Despite further research on the subject, results are still mixed across the
literature. Mani et al. (2018) investigated the effect of ALAD polymorphisms on 561 occupationally Pb-
exposed and 317 nonoccupationally Pb-exposed subjects in India. The mean BLL levels for the
occupationally exposed group were 57.69 ± 29.1 (ALAD 1-2/2-2) and 53.97 ± 28.62 (ig/dL (ALAD 1-1),
whereas for the nonoccupationally exposed group, BLLs were 3.83 ± 2.65 (ALAD 1-2/2-2) and
3.25 ± 2.26 (ig/dL (ALAD 1-1). Sobin et al. (2011) investigated the association of BLLs and ALAD
polymorphisms in 306 minority children in Texas. Heterozygous boys with ALAD-2 present had a mean
BLL of 3.5 (ig/dL, whereas those without had a mean of 2.7 (ig/dL. Heterozygous girls with ALAD-2
present had a mean BLL of 2.6 (ig/dL, whereas those without had a mean of 2.7 (ig/dL. Kavaalti et al.
(2016) studied placental Pb levels in a small sample of 97 pregnant women in Turkey and found those
with ALAD 1-1, ALAD 1-2, and ALAD 2-2 polymorphisms had median values of 7.54 |ig/kg.
11.78 |ig/kg. and 18.53 |ig/kg. respectively. In another small study of 81 brain tumor patients in Egypt,
mean BLLs for those with the presence of only ALAD-1 and those with an ALAD-2 allele were found to
be 25.93 (ig/dL ± 12.73 and 34.39 (ig/dL ± 17.87, respectively. In contrast, Warrington et al. (2015).
using Australian and U.K. cohorts, found no statistically significant association of BLLs with ALAD 1-2.
Lerover et al. (2013) studied ALAD polymorphism and BLLs in 204 French men, finding no statistically
significant difference between those with ALAD 1-1, ALAD 1-2, or ALAD 2-2 polymorphisms.
Saturable binding to RBC proteins contributes to an increase in the plasma/blood Pb ratio with
increasing PbB and curvature to the blood Pb-plasma Pb relationship (Rentschler et al.. 2012; Kang et al..
2009; Jin et al.. 2008; Barbosa et al.. 2006a; Smith et al.. 2002; Manton et al.. 2001; Bergdahl et al.. 1999;
Bergdahl et al.. 1998; Bergdahl et al.. 1997b; deSilva. 1981). An example of this is shown in Figure 2-3.
Saturable binding of Pb to RBC proteins has several important consequences. As blood Pb increases and
the higher affinity binding sites for Pb in RBCs become saturated, a larger fraction of the blood Pb is
available in plasma to distribute to the brain and other Pb-responsive tissues. This change in distribution
2-57
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of Pb contributes to a curvature in the relationship between Pb intake (at constant absorption fraction) and
blood Pb concentration. Plasma Pb also exhibits faster kinetics. Following exposures of five adults that
resulted in relatively high blood Pb concentrations (56-110 (.ig/dL). the initial (fast-phase) elimination
half-time for plasma Pb (38 ± 20 [SD] days) was approximately half that of blood (81 ± 25 days)
(Rentschler et al.. 2012).
Typically, at blood Pb concentrations <100 (ig/dL, only a small fraction (<1%) of blood Pb is
found in plasma (Marcus. 1985; Manton and Cook. 1984; deSilva. 1981). However, as previously noted,
plasma Pb may be the more biologically labile and toxicologically effective fraction of the circulating Pb.
Approximately 40%-75% of Pb in the plasma is bound to proteins, of which albumin appears to be the
dominant ligand (Al-Modhefer et al.. 1991; Ong and Lee. 1980). Pb in serum that is not bound to protein
exists largely as complexes with low molecular weight sulfhydryl compounds (e.g., cysteine,
homocysteine) and other ligands (Al-Modhefer et al.. 1991).
2.0
1.5
T3
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£ i.o
CO
E
V)
ro
o- 0.5
0.0
0
oAdults •Children
•
r
20 40 60
Blood Pb (|jg/dL)
80
100
Source: Adapted with permission of Elsevier Publishing and the Finland Institute of Occupational Health, Beradahl et al. CI9991:
Beradahl et al. d997b1.
Figure 2-3 Plot of blood and plasma Pb concentrations measured in adults
and children.
As shown in Figure 2-3, the limited binding capacity of Pb-binding proteins in RBCs produces a
curvilinear relationship between blood and plasma Pb concentration. The limited binding capacity of RBC
binding proteins also confers, or at least contributes, to a curvilinear relationship between Pb intake and
blood Pb concentration. A curvilinear relationship between Pb intake and blood Pb concentration has
2-58
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been observed in children (Sherlock and Ouinn. 1986; Lacev et al.. 1985; Ryu et al.. 1983). Data from
Sherlock and Quinn (1986) are illustrated in Figure 2-4; although the blood Pb is limited to >13 (ig/dL,
the relationship becomes approximately linear at relatively low daily Pb intakes (i.e., <50 (ig/day) and
blood Pb concentrations <22 (ig/dL.
Pb Intake (|jg/day)
Data represent mean and standard errors for intake; the line is the regression model (blood Pb = 3.9 + 2.43 (Pb intake [|jg/week]1'3).
Source: Adapted with permission of Taylor & Francis Publishing, Sherlock and Quinn CI 9861.
Figure 2-4 Relationship between Pb intake and blood Pb concentration in
infants (n = 105, age 13 weeks, formula fed).
Figure 2-5 shows the predicted relationship between quasi-steady state blood and plasma Pb
concentrations in a 4-year-old child using the International Commission on Radiological Protection
(ICRP) model (Pounds and Lcggctt. 1998; Leggett. 1993). The ICRP model is a mechanistic model of Pb
biokinetics that consists of a systemic biokinetics model (Leggett. 1993) and absorption factors for
inhaled Pb (ICRP. 1995) (see Section 2.6 for a brief description). The abrupt inflection point that occurs
at approximately 25 (ig/dL blood Pb is an artifact of the numerical approach to simulate the saturation of
binding using discontinuous first-order rate constants for uptake and exit of Pb from the RBC. A
continuous function of binding sites and affinity, using empirical estimates of both parameters, yield a
similar but continuous curvature in the relationship (Bergdahl et al.. 1998; O'Flahertv. 1995).
Nevertheless, either approach predicts an approximately linear relationship at blood Pb concentrations
below about 25 (ig/dL, which, in this model, corresponds to an intake of about 100 (ig/day (absorption
rate ~ 30 (ig/day) (upper panel). An important consequence of the limited Pb-binding capacity of RBC
2-59
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proteins is the plasma Pb concentration will continue to grow at a linear rate above the saturation point for
RBC protein binding. One implication of limited RBC binding capacity is a larger fraction of the Pb in
blood will be "free" in plasma and available to distribute to the brain and other tissues as blood Pb
increases. This process could potentially contribute to nonlinearity in dose-response relationships in
studies in which blood Pb is the used as the internal dose metric.
2-60
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0.60
0.50 -
1) 0.40 H
3
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ra
ra 0.20 -
0.10 -I
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-i-
10
20 30 40 50
Blood Pb (|ig/dL)
60
50 -
T3
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30 -
.Q
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O
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100 200 300
Intake (tig/day)
1.0
0.8
0.6
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0.4 £
era
0.2
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400
Note: Model simulations are for a 4-year-old having from birth a constant Pb intake of between 1 and 400 |jg/day. Simulation based
on ICRP Pb biokinetics model CLegqett. 19931 with tissue and compartment masses and volumes based on equations and
parameters from O'Flaherty's studies (O'Flahertv. 1995. 1993).
Figure 2-5 Simulation of quasi-steady state blood and plasma Pb
concentrations in a child (age 4 years) associated with varying Pb
ingestion rates.
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Studies conducted in swine provide additional evidence in support of RBC binding kinetics
influencing distribution of Pb to tissues. In these studies, the relationship between the ingested dose of Pb
and tissue Pb concentrations (e.g., liver, kidney, bone) was linear, whereas the relationship between dose
and blood Pb was curvilinear with the slope decreasing as the dose increased (Casteel et al.. 2006).
Saturable binding of Pb to RBC proteins also contributes to a curvilinear relationship between blood Pb
and both plasma Pb and urinary Pb, whereas Pb in plasma and urine are linearly related (Bcrgdahl et al..
1997b).
2.2.2.2 Bone
The dominant compartment for Pb in the body is in bone. In human adults, more than 90% of the
total body burden of Pb is found in the bones, whereas bone Pb accounts for just under 60% of the body
burden in infants less than a year old and just over 70% of the body burden in older children (Barry.
1975). Bone is composed of two main types, cortical (or compact) and trabecular (or spongy or
cancellous). The proportion of cortical to trabecular bone in the human body varies by age but is about 80
to 20% in adults (O'Flahertv. 1998; Leggett. 1993; ICRP. 1973). It should be recognized that cortical and
trabecular bone coexist within the same bone. For example, the tibia is generally considered a cortical
bone with less than 1% trabecular bone at its midshaft but is 55%-75% trabecular bone toward the ends
of the bone [see paragraph 38 of ICRP (1996)1. In totality, the tibia is 74%-83% cortical and 17%-26%
trabecular. Compact cortical bone is found along the shaft (diaphysis) of long bones, whereas the spongy,
more highly perfused trabecular bone is found toward the ends (metaphysis) of the bones where growth is
occurring and further out (epiphysis) toward the ends of the bones (ICRP. 2002a).
Pb distribution in bone includes uptake into cells that populate bone (e.g., osteoblasts, osteoclasts,
osteocytes) and exchanges with proteins and minerals in the extracellular matrix (Pounds et al.. 1991). Pb
forms highly stable complexes with phosphate and can replace calcium in the calcium-phosphate salt,
hydroxyapatite, which comprises the primary crystalline matrix of bone (Meirer et al.. 2011; Bres et al..
1986; Mivake et al.. 1986; Verbeeck et al.. 1981). Several intracellular kinetic pools of Pb have been
described in isolated cultures of osteoblasts and osteoclasts, which appear to reflect physiological
compartmentalization within the cell, including membranes, mitochondria, soluble intracellular binding
proteins, mineralized Pb (i.e., hydroxyapatite) and inclusion bodies (Long et al.. 1990; Pounds and Rosen.
1986; Rosen. 1983). Approximately 70%-80% of Pb taken up into isolated primary cultures of
osteoblasts or osteocytes is associated with mitochondria and mineralized Pb (Pounds et al.. 1991).
The composition of bones changes with age (ICRP. 2002a). In infants, compact cortical bone is
highly vascular with a large portion of bone surfaces showing formation (calcification) and reabsorption.
By adolescence, the cortical bone is more stable in structure and uniform in appearance. By later
adulthood, cortical bone begins to become porous. For trabecular bone, there appears to be a rapid
increase during infancy that may continue more gradually through childhood followed by a slow decline
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thereafter. There may be changes in trabecular bone after bone growth has ceased in response to
mechanical stress on the bone. With the changes in bone composition, the density of hydrated bone
increases from birth to adulthood, but then decreases beyond about 40 years of age.
Pb accumulates in bone regions having the most active calcification at the time of exposure. In the
2006 Pb AQCD (U.S. EPA, 2006) and 2013 Pb ISA (U.S. EPA, 2013), Pb accumulation is thought to
occur predominantly in cortical bone during childhood and in both cortical and trabecular bone in
adulthood. However, considering the changes in bone composition early in life, a rigid dichotomy
between accumulation of Pb in trabecular versus cortical bone during childhood is complicated. With
continued exposure, Pb concentrations in bone may increase with age throughout the lifetime beginning
in childhood, indicative of a relatively slow turnover of Pb in adult bone (Park et al„ 2009; Barry and
Connolly, 1981; Barry, 1975; Gross et al., 1975; Schroeder and Tipton, 1968). The cortical and trabecular
bones have distinct rates of turnover and Pb release, which is about 1.5-1.7 times greater in adults for
trabecular than cortical bone in terms of both volume and grams calcium per day [see Table 20 of ICRP
(1996)1.
A high bone formation rate in early childhood results in the rapid uptake of circulating Pb into
mineralizing bone; however, bone Pb is also recycled to other tissue compartments, back to bone, or
excreted in accordance with a high bone resorption rate (O'Flaherty, 1995). Thus, most (60%-65%) of the
Pb acquired early in life is not permanently fixed in the bone (O'Flaherty, 1995; Lcggctt. 1993; ICRP,
1973). However, some Pb accumulated in bone may persist into later life. McNeill et al. (2000) compared
tibia Pb levels and cumulative blood Pb indices in a population of 19- to 29-year-olds who had been
highly exposed to Pb in childhood from the Bunker Hill, ID smelter; they concluded Pb from exposure in
early childhood had persisted in the bone matrix until adulthood.
Additional discussion of the Pb in bone and its mobilization are provided in other Sections of this
chapter. Maternal mobilization of Pb from the bone to the fetus is discussed in Section 2.2.2.4. The
relationship between Pb in blood and bone is discussed in Section 2.3.5.
2.2.2.3 Soft Tissues
Most of the Pb in soft tissue is in the liver and kidney (Gerhardsson et al.. 1995; Oldereid et al..
1993; Gerhardsson et al.. 1986; Barry. 1981; Barry. 1975; Gross et al.. 1975). Presumably, the Pb in these
soft tissues (i.e., kidney, liver, and brain) exists predominantly bound to protein. High-affinity cytosolic
Pb-binding proteins have been identified in rat kidney and brain (DuVal and Fowler. 1989; Fowler.
1989). The Pb-binding proteins in rats are cleavage products of a2\i globulin, a member of the protein
superfamily known as retinol-binding proteins that are generally observed only in male rats (Fowler and
DuVal. 1991). Other high-affinity Pb-binding proteins (Kd -14 nM) have been isolated in human kidney,
two of which have been identified as a 5 kDa peptide, thymosin 4 and a 9 kDa peptide, acyl-CoA binding
protein (Smith et al.. 1998). Pb also binds to metallothionein but does not appear to be a significant
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inducer of the protein in comparison with the inducers Cd and Zn (Waalkcs and Klaassen. 1985; Eaton et
al.. 1980).
The liver and kidneys rapidly accumulate systemic Pb (ti/2 = 0.21 and 0.41 hours, respectively),
which amounts to 10%—15% and 15%-20% of intravenously injected Pb, respectively (Leggett. 1993). A
linear relationship in dose-tissue Pb concentrations for kidney and liver has been demonstrated in swine,
dogs, and rats (Smith et al.. 2008; Casteel et al.. 2006; Casteel et al.. 1997; Azar et al.. 1973). In contrast
to bone, which accumulates Pb with continued exposure in adulthood, concentrations in soft tissues
(e.g., liver and kidney) are relatively constant in adults (Treble and Thompson. 1997; Barry. 1975).
reflecting a faster turnover of Pb in soft tissue relative to bone.
2.2.2.4 Fetus
Evidence for maternal-to-fetal transfer of Pb in humans is derived from cord blood Pb to maternal
blood Pb ratios (i.e., cord blood Pb concentration divided by mother's blood Pb concentration). Group
mean ratios range from about 0.7 to 1.0 at the time of delivery for mean maternal BLLs ranging from 1.7
to 8.6 (ig/dL (Rollin et al.. 2017; Amaral et al.. 2010; Kordas et al.. 2009; Patel and Prabhu. 2009;
Carbone et al.. 1998; Gover. 1990; Graziano et al.. 1990). The relationship for individual mother-child
pairs is variable but well correlated (Pearson r = 0.79); in a predominantly young, low-income, urban
population (n = 159), factors associated with higher cord BLL compared with maternal BLL included
maternal elevated blood pressure and alcohol consumption, whereas factors associated with relatively
lower ratios of cord blood Pb to maternal blood Pb included maternal increased hemoglobin levels and
sickle cell trait (Harville et al.. 2005). Calcium intake and physical activity were not associated with
differences between cord blood Pb and maternal blood Pb. Consistent with other studies, the ratio of mean
cord blood Pb (1.64 (ig/dL) to mean maternal blood Pb (1.93 (ig/dL) was 0.85. The similarity of isotopic
ratios in maternal blood and in blood and urine of newly born infants provides further evidence for
placental transfer of Pb to the fetus (Gulson et al.. 1999).
Transplacental transfer of Pb may be facilitated by an increase in the plasma/blood Pb
concentration ratio during pregnancy (Montenegro et al.. 2008; Lamadrid-Figueroa et al.. 2006).
Maternal-to-fetal transfer of Pb appears to be related partly to the mobilization of Pb from the maternal
skeleton. Evidence for transfer of maternal bone Pb to the fetus has been provided by stable Pb isotope
studies in cynomolgus monkeys exposed during pregnancy. Approximately 7%-39% of the maternal Pb
burden transferred to the fetus was derived from the maternal skeleton, with the remainder derived from
contemporaneous exposure (O'Flahertv. 1998; Franklin et al.. 1997). The upper value in the range (39%)
represented the one monkey with historical Pb exposure from a brief 4-month exposure period in 1990
with 2ll4Pb acetate trihydrate (nominally 1,500 |ig Pb/kg/day) but received only small amounts of
environmental Pb exposure during pregnancy; for the monkeys that received high doses of Pb during
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pregnancy (1,500 |ig Pb/kg/day; 7 days/week), the range was lower (7%-25%) (O'Flahcrtv. 1998;
Franklin et al.. 1997).
2.2.2.5 Organic Pb
Information on the distribution of Pb in humans following exposures to organic Pb is extremely
limited. However, as reported in the 2006 Pb AQCD (U.S. EPA, 2006), the available evidence
demonstrates near complete absorption following inhalation of tetraalkyl Pb vapor and subsequent
transformation to trialkyl Pb metabolites. One hour following brief inhalation exposures to 2ll3Pb
tetraethyl or tetramethyl Pb (1 mg/m3), -50% of the 2ll3Pb body burden was associated with liver and 5%
with kidney; the remaining 2ll3Pb was widely distributed throughout the body (Heard et al„ 1979). The
kinetics of 2ll3Pb in blood showed an initial declining phase during the first 4 hours (tetramethyl Pb) or
10 hours (tetraethyl Pb) after the exposure, followed by a reappearance of radioactivity back into the
blood after -20 hours. The high level of radioactivity initially in the plasma indicates the presence of
tetraalkyl/trialkyl Pb. The subsequent rise in blood radioactivity, however, probably represents water-
soluble inorganic Pb and trialkyl and dialkyl Pb compounds that were formed from the metabolic
conversion of the volatile parent compounds (Heard et al„ 1979).
Alkyl Pb compounds undergo oxidative dealkylation catalyzed by cytochrome P450 in the liver
and, possibly, other tissues. Trialkyl Pb metabolites have been found in the liver, kidney, and brain
following exposure to the tetraalkyl compounds in workers (Bolanowska et al., 1967); these metabolites
have also been detected in brain tissue of nonoccupational subjects (Nielsen et al., 1978).
2.2.3 Elimination
The rapid phase (30 to 40 days) of Pb excretion in adults accounts for a varied fraction of
absorbed Pb (Chamberlain et al.. 1978; Rabinowitz et al.. 1976; Kehoe. 1961a. b). The fraction of
absorbed Pb that is rapidly eliminated generally decreases with increasing exposure duration. This rapid
phase of Pb excretion is followed by slower phases of Pb clearance from soft tissues and bone. Due to the
long half-life of Pb in bone, it can serve to maintain BLLs long after external exposure has ceased.
Absorbed Pb is excreted primarily in urine and feces, with sweat, saliva, hair, nails, and breast milk being
minor routes of excretion (Kehoe. 1987; Chamberlain et al.. 1978; Rabinowitz et al.. 1976; Griffin et al..
1975; Hursh et al.. 1969; Hursh and Suomela. 1968).
Approximately 30% of intravenously injected Pb in humans (40%-50% in beagles and baboons)
is excreted via urine and feces during the first 20 days following administration (Leggett. 1993). The
kinetics of urinary excretion following a single dose of Pb is similar to that of blood (Chamberlain et al..
1978). likely due to the fact that Pb in urine derives largely from Pb in plasma. Evidence for this is the
observation that urinary Pb excretion is strongly correlated with the rate of glomerular filtration of Pb
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(Araki et al.. 1986) and plasma Pb concentration (Rcntschlcr et al.. 2012; Bergdahl et al.. 1997b)
(i.e., glomerular filtration rate x plasma Pb concentration), and both relationships are linear. While the
relationship between urinary Pb excretion and plasma Pb concentration is linear, the plasma Pb
relationship to blood Pb concentration is curvilinear (as described in Section 2.2.2.1 and demonstrated in
Figure 2-3). This relationship contributes to an increase in the renal clearance of Pb from blood with
increasing blood Pb concentrations (Chamberlain. 1983). Similarly, a linear relationship between plasma
Pb concentration and urinary excretion rate predicts a linear relationship between Pb intake (at constant
absorption fraction) and urinary Pb excretion rate, whereas the relationship with blood Pb concentration
would be expected to be curvilinear (Section 2.3.6).
Estimates of urinary filtration of Pb from plasma range from 13 to 22 L/day, with a mean of
18 L/day (Araki et al.. 1986; Manton and Cook. 1984; Manton and Mallov. 1983; Chamberlain et al..
1978). which corresponds to half-time for transfer of Pb from plasma to urine of 0.10 to 0.16 days for a
70 kg adult who has a plasma volume of ~3 L. The rate of urinary excretion of Pb was less than the rate of
glomerular filtration of ultrafilterable Pb, suggesting urinary Pb is the result of incomplete renal tubular
reabsorption of Pb in the glomerular filtrate (Araki et al.. 1986); however, net tubular secretion of Pb has
been demonstrated in animals (Victerv et al.. 1979; Vander et al.. 1977). On the other hand, estimates of
blood-to-urine clearance range from 0.03 to 0.3 L/day with a mean of 0.18 L/day (Diamond. 1992; Araki
et al.. 1990; Bergeretal.. 1990; Koster et al.. 1989; Manton and Mallov. 1983; Ryu et al.. 1983;
Chamberlain et al.. 1978; Rabinowitz et al.. 1973). consistent with a plasma Pb to blood Pb concentration
ratio of-0.005-0.01 L/day (U.S. EPA. 2003a).
More recently, Diamond et al. (2019) estimated blood-to-urine clearance in adolescents (12 to
<20 years; n = 1,269) and adults (20 to 80 years; n = 6,356) using paired blood Pb, urine Pb, serum
creatinine, and urine creatinine concentration data in individual subjects from 2009 to 2016 NHANES
data. The median (5th, 95th percentile range) blood-to-urine clearance rates were 0.043 (0.008,
0.132) L/day in adolescents and 0.040 (0.009, 0.118) L/day in adults. Linear regression, including age,
gender (NHANES variable was self-identified sex), body weight for adults, body height for adolescents,
and serum creatinine clearance (a metric of the glomerular filtration rate, GFR) explained 67%-68% of
the variability in blood-to-urine clearance. Serum creatinine clearance (i.e., GFR) accounted for 95%-
98% of the explained variance in blood-to-urine clearance. On the basis of the above differences, urinary
excretion of Pb can be expected to reflect the concentration of Pb in plasma and variables that affect
delivery of Pb from plasma to urine (e.g., glomerular filtration and other transfer processes in the kidney).
Ho et al. (2022) investigated an index of blood-to-urine clearance in adolescents (12 to 18 years;
1,542 males and 1,383 females) using paired blood Pb, urine Pb, serum creatinine, and urine creatinine
concentration data of individual subjects from 1999-2012 NHANES data. The authors normalized urine
Pb for dilution by dividing by urine creatinine. The authors observed the ratio of normalized urine Pb to
blood Pb was 30% lower in males than females. On the basis of this observation, the authors suggested
differences in renal elimination contributed to a greater body burden (as indicated by blood Pb) in males
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relative to females. However, the normalized urine Pb to blood Pb ratio used by the authors is not a
measure of urinary Pb elimination. Urinary Pb elimination requires using a measure of total urine flow as
conducted by Diamond et al. (2019). who found the ratio of urinary Pb elimination rate to blood Pb was
very similar between males and females.
The value for fecal:urinary excretion ratio (-0.5) was observed during days 214 following
intravenous injection of Pb in humans (Chamberlain et al.. 1978; Booker et al.. 1969; Hurshetal.. 1969).
This ratio is slightly higher (0.7 to 0.8) with inhalation of submicron Pb-bearing PM due to ciliary
clearance and subsequent ingestion. The transfer of Pb from blood plasma to the small intestine by biliary
secretion in the liver is rapid (adult ti/2 = 10 days) and accounts for 70% of the total plasma clearance
(O'Flahertv. 1995).
Organic Pb absorbed after inhalation of tetraethyl and tetramethyl Pb is excreted in exhaled air,
urine, and feces (Heard etal.. 1979). Fecal:urinary excretion ratios were 1.8 following exposure to
tetraethyl Pb and 1.0 following exposure to tetramethyl Pb (Heard et al.. 1979). Occupational monitoring
studies of workers exposed to tetraethyl Pb showed it is excreted in the urine as diethyl Pb, ethyl Pb, and
inorganic Pb (Yural and Duvdu. 1995; Zhang et al.. 1994; Turlakiewicz and Chmielnicka. 1985).
2.3 Pb Biomarkers
The 2013 Pb ISA (U.S. EPA. 2013) contains background information on Pb in various
biomarkers and their relationships. This section explores recent advances in the biological measurements
of Pb that act as indicators of exposure or body burden and the relationships between those biomarkers,
including bone and blood Pb. Although the following Sections look at Pb in different biomarkers
individually, body burden can be represented by multiple biomarkers at the same time. Levin-Schwartz et
al. (2020) proposed the concept of a multimedia biomarker (MMB) for Pb. In their study, they developed
a weighting of multiple biomarkers, including blood Pb, to represent body burden. They found blood Pb
and the developed MMB best correlated with IQ scores for 251 Italian adolescents.
2.3.1 Bone-Pb Measurements
Because mineralized tissues within the body act as long-term Pb storage sites with a half-life
measured in decades, measurement of Pb within these tissues is important to understand overall body
burden. Bone measurements of Pb are conducted through a variety of methods that can be invasive or
noninvasive. The 2013 Pb ISA (U.S. EPA. 2013) contains a comprehensive list of invasive methods that
measure Pb concentration in excised bone, including flame AAS and anodic stripping voltammetry
(ASV). Noninvasive in vivo measurements can be done using XRF. As the 2013 Pb ISA (U.S. EPA.
2013) noted, the rise in the popularity of XRF as a measurement tool for bone Pb has eclipsed other
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methods because of its ease of use. K-shell X-ray fluorescence (K-XRF) has been used widely to conduct
in vivo measurements of both trabecular and cortical bone (Specht et al.. 2016).
XRF is now incorporated into portable technologies (Nie et al.. 2011a). Zhang et al. (2021)
evaluated a portable XRF device against a traditional K-XRF instrument using the mid-tibia bone in 71
people of three Indiana communities. The correlation between the portable XRF and K-XRF instruments
for all participants was r = 0.48 (95% CI: 0.27, 0.64). However, correlation was much higher r = 0.78
(95% CI: 0.61, 0.87) for those with minimal soft tissue thickness (>5 mm). Portable XRF works most
accurately on bones with minimum tissue thickness such as the skull and tibia (Specht et al.. 2019a).
Given the shallow penetration depth (0.2 mm) of portable XRF and the fact that all bone is covered in a
cortical shell, it is likely that any portable XRF measurements are of cortical bone.
New developments are allowing for Pb to be spatially resolved. Pemmer et al. (2013) found,
using XRF spatial mapping of 14 human bone samples from individuals with osteoporotic femoral neck
fractures, that levels of Pb accumulated in the cement lines of samples was roughly two times more than
the surrounding bone matrix. Specht et al. (2019a) used portable XRF along the skulls and tibias of 31
cadavers, finding no real change in Pb levels, matching previous studies of the skull and tibia.
2.3.2 Blood-Pb Measurements
The 2013 Pb ISA (U.S. EPA. 2013) details common methods and their limitations for screening
Pb in blood, including AAS, graphite furnace atomic absorption spectrometry, ASV, ICP-AES, and ICP-
MS. Blood measurements can be taken through venous blood samples or capillary blood samples.
Capillary blood samples are commonly collected due to their ease of collection (i.e., a finger prick) versus
venipuncture for venous blood samples. Point-of-care instruments using ASV offer low-cost, "in office"
results within minutes (ACCLPP. 2013). Anderson et al. (2007) examined false positive capillary blood
samples in 0- to 5-year-old children between 2002 and 2003 in Maine. Defining a false positive as a
capillary BLL >10 (ig/dL with a confirmatory venous BLL <10 (ig/dL, they found a 73% false positive
rate. False positive capillary samples were most frequent for BLL between 10 and 14 (ig/dL (i.e., just in
excess of the former CDC blood Pb action level). Using 2011-2017 Minnesota data for 0- to 6-year-old
children, Wang et al. (2019) reported 60% false positives, defined as having a capillary BLL >5 (ig/dL
followed by a venous BLL <5 (ig/dL. False positive capillary samples were most frequent for BLL
between 5 and 6.9 (ig/dL (i.e., just over the CDC's 2012 blood Pb reference value, BLRV). False positive
capillary BLLs are due to a positive bias in capillary sample measurement and contamination of the
fingertips where samples were collected (Wang et al.. 2019; Anderson et al.. 2007).
There are challenges to measuring BLLs at low values, especially as average blood Pb
concentrations become lower as a result of reductions in exposure. At lower BLLs, contamination of
equipment also becomes a larger issue. Pb contamination can occur in laboratory reagents and supplies
and during sample collection. Laboratories have had to update equipment to measure at lower limits of
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detection from flame absorption spectroscopy in the 1970s to newer methods of ICP-MS analysis used
today. Caldwell et al. (2017) presents data from the Lead and Multi-Element Proficiency program, which
evaluated the performance of BLL measurements in approximately 180 laboratories between 2011 and
2015. Although the study found most U.S. laboratories can measure BLLs at ±2 (ig/dL (<20 |ig/dL). the
authors noted the current acceptability criteria for BLL measurements is ±4 (ig/dL or ±10%, whichever is
greater. Measurement precision of laboratories was quantified in terms of an RSD, which increases with
decreasing BLL. For four BLL samples sent to 50 labs (205 total labs) and consensus mean BLLs ranging
from 1.1 to 1.5 (ig/dL (average 1.3 |ig/dL). the RSD ranged from 37% to 70% (average 49%). For five
BLL samples sent to 50 labs (247 total labs) and consensus mean BLLs ranging from 4.15 to 5.15 (ig/dL
(average 4.7 |ig/dL). the RSD ranged from 10% to 19% (average 17%). On the basis of these data and
assuming a linear trend, RSD is estimated to be about 28% and 14% for 3.5 and 5 (ig/dL, respectively.
For a single blood Pb measurement from a child, the 95% CI are 1.6-5.4 (ig/dL and 3.7-6.3 (ig/dL for the
actual BLLs of 3.5 and 5 (ig/dL. While a measured value of 5 (ig/dL showed a child's BLL was over
3.5 (ig/dL, a measured BLL of 3.5 (ig/dL should not necessarily be clinically interpreted as showing the
child has a BLL of <5 (ig/dL. In a subsequent CDC study, Caldwell et al. (2019) reported laboratory
precision ranged from 0.26 (ig/dL for ICP-MS to 1.50 (ig/dL for ASV.
Hague et al. (2021) proposed a method for measurement of Pb in the archived clotted erythrocyte
fraction of whole blood. A Pearson correlation coefficient of 0.90 and 0.89 were found for acid digestion
and alkaline dilution, respectively.
2.3.3 Urine-Pb Measurements
The 2013 Pb ISA (U.S. EPA. 2013) summarizes issues related to using urine Pb as a biomarker.
Briefly, the concentration of Pb in urine is a function of urinary Pb excretion and flow rate. Urine samples
can be collected as timed or untimed samples, with untimed samples needing correction to account for
variation in urine flow, which can vary by a factor of more than 10. Urine-Pb concentration measurements
provide little reliable information about exposure or body burden unless they can be adjusted to account
for unmeasured variability in flow rate. Urine-Pb concentration reflects concentration of Pb in blood,
representing both recent and past exposures to Pb, and thus cannot distinguish between a long-term low
level of exposure or a higher acute exposure. The literature search and screening for this appendix did not
capture any significant new advancements in methodology for urine-Pb measurements. A discussion of
urinary Pb elimination is provided in Section 2.2.3.
2.3.4 Pb in Other Biomarkers
The 2006 Pb AQCD (U.S. EPA, 2006) contains detailed discussion on using Pb biomarkers
other than blood Pb or bone Pb as indicators of exposure. The 2013 Pb ISA (U.S. EPA, 2013) contains
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additional summaries of what is known regarding the use of teeth, hair, saliva, and serum 8-
aminolevulinic acid (S-ALA) and ALAD as biomarkers of Pb exposure. These other biomarkers have not
been established to the same extent as blood and bone Pb. Below are summaries of recent literature
containing information on advances in methodology for measurement of these biomarkers.
2.3.4.1 Teeth
As discussed in the 2013 Pb ISA (U.S. EPA. 2013). researchers have advocated use of sections of
the enamel and dentine to obtain more information on Pb exposure rather than using the whole tooth. Two
popular analytical techniques, among others, are laser ablation-inductively coupled plasma-mass
spectrometry (LA-ICP-MS) and microbeam synchrotron radiation X-ray fluorescence (fi-SRXRF). Both
of these techniques have been proposed for measurement of sections of teeth rather than the whole tooth
to understand timing of exposure as the tooth develops. Teeth are composed of several tissues formed pre-
and postnatal. Therefore, if a child's Pb exposure during the years of tooth formation varied widely,
different amounts of Pb would be deposited at different rates (Rabinow itz et al.. 1993). The neonatal line
formed in deciduous teeth during birth can be used to distinguish between prenatal and postnatal dentine
and enamel (Hodgson et al.. 2015). Shepherd et al. (2012) and Shepherd et al. (2016) used LA-ICP-MS
across two small samples of deciduous teeth to reconstruct histories of exposure.
Aroraet al. (2014) proposed measuring Pb in prenatal, postnatal, and secondary dentine of 34
incisors, 25 canines, and 26 molars naturally shed from children using LA-ICP-MS. They found strong
association between birth dentine Pb and maternal cord blood Pb with a weaker association as the child
aged. Johnston et al. (2019) also used this technique to assess the correlation of prenatal tooth Pb and
postnatal tooth Pb with surrounding soil Pb levels in 43 child subjects in Los Angeles, CA. After
adjusting for maternal education and batch, positive associations were observed between teeth Pb
concentration per 100 mg/L increase in soil Pb concentration for both prenatal teeth Pb (statistically
significant) and postnatal teeth Pb (p = 0.056). Wang et al. (2017b) applied (J.-SRXRF to one incisor and
two molars. The authors were able to successfully resolve Pb concentrations at the micrometer scale.
2.3.4.2 Hair
The 2006 Pb AQCD (U.S. EPA, 2006) discusses the applications, methodological limitations
(e.g., external contamination), and lack of empirical basis for using hair Pb as a biomarker of Pb
exposure. The 2013 Pb ISA (U.S. EPA, 2013) summarizes this information. Although several studies
have used hair as a biomarker for Pb exposure since 2011, there have been no major methodological
advancements, and there are still major limitations present (Skroder et al„ 2017). Hair Pb measurements
may be contaminated at the surface by environmental Pb or artificial hair treatments. They are also a poor
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predictor of blood Pb (U.S. EPA. 2013). Pb concentrations have been found to vary along the hair shaft
(Jursa et al.. 2018).
2.3.4.3 Saliva
Sampling salivary Pb is an attractive alternative to blood Pb sampling because of its ability to be
noninvasive. The 2013 Pb ISA (U.S. EPA. 2013) summarizes earlier literature on salivary Pb
measurements. It indicates older reports of salivary Pb showed strong correlation between blood Pb and
salivary Pb but reports between 2006 and 2011 showed weak or inconsistent associations. Both Barbosa
et al. (2006b) and Nriagu et al. (2006) found significant but weak associations between blood Pb and
salivary Pb in adults from two different populations. These differences in outcomes may be a result of
exposure history, dental health, and/or the methods for determining Pb in saliva.
Staff et al. (2014). when collecting saliva and whole blood from 105 U.K. workers, noted that
refrigerated blank saliva run through the saliva collection device resulted in significant Pb contamination
from the device itself that was also highly variable. Additionally, their review of the literature found the
correlation between salivary Pb and blood Pb was much stronger at higher Pb levels than low exposure
levels, with their own study having a Pearson's r of 0.457 between log(salivary Pb) and log(blood Pb).
When testing 407 oral fluid samples of children aged 6 months to 5 years, Gardner et al. (2016) found a
Pearson's r of 0.687 between blood Pb and salivary Pb samples. Given currently available data and lack
of uniform testing methods and conditions, it is unclear whether salivary Pb can be a more reliable testing
method than blood Pb measurements.
2.3.4.4 Serum 6-ALA and ALAD
The 2013 Pb ISA (U.S. EPA. 2013) concluded blood ALAD activity and serum 8-ALA could
potentially be used as biomarkers for Pb exposure. Inhibition of erythrocyte ALAD by Pb results in a rise
of the ALAD substrate S-ALA in plasma. Huang et al. (2020) investigated the threshold of ALAD activity
reduced by Pb exposure by using BLL and polymorphism data from 121 Pb workers and 117 nonexposed
workers in Taiwan. Using a generalized additive model and multiple regressions, the authors found BLLs
above 10 (ig/dL resulted in significantly inhibited ALAD enzyme activity. On the basis of the different
ranges of BLLs studied, the authors recommend a range of 5-10 (ig/dL as an inflection point for declining
ALAD activity among adults. This is similar to La-Llave-Leon et al. (2017). who found, among 633
pregnant women in Mexico, that ALAD activity was reduced for BLLs between 2.2 and 10 (ig/dL.
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2.3.5
Relationship between Pb in Blood and Pb in Bone
The kinetics of elimination of Pb from the body reflects the existence of multiple pools of Pb in
the body. The dominant washout phase of Pb from the blood, exhibited shortly after a change in exposure
occurs, has a half-life of -20-30 days (Leggett. 1993: Rabinowitz et al.. 1976) in adults. Studies of a
limited number of adults (four individuals with hip or knee replacement, a married couple, and 10 female
Australian immigrants) in which the Pb exposure was from historical environmental sources
(i.e., minimal current Pb exposure relative to past Pb exposure) have found bone Pb stores can contribute
as much as 40%-70% to blood Pb (Smith et al.. 1996; Gulson et al.. 1995; Manton. 1985). Bone Pb
burdens in adults are slowly lost by diffusion (heteroionic exchange) as well as by bone resorption
(O'Flahertv. 1995). Half-times for the release of Pb in bone are dependent on age and intensity of
exposure. Bone compartments are much more labile in infants and children than in adults as reflected by
half-times for movement of Pb from bone into the plasma (e.g., cortical ti/2 = 0.23 years at birth, 1.2 years
at 5 years of age, 3.7 years at 15 years of age, and 23 years in adults; trabecular ti/2 = 0.23 years at birth,
1.0 years at 5 years of age, 2.0 years at 15 years of age, and 3.9 years in adults) (Leggett. 1993). Slow
transfer rates for the movement of Pb from nonexchangeable bone pools to plasma are the dominant
transfer process determining long-term accumulation and elimination of bone Pb burden.
Pb transferred from bone and other body compartments to plasma, as well as newly absorbed Pb
from the GI and respiratory tracts, is, in part, transferred to bone surfaces. The exchange of Pb from
plasma to the bone surface is a rapid process. On the basis of Leggett (1993). the half-time for movement
of Pb from plasma to trabecular bone surface is 11 minutes in an adult and 17 minutes in a 1-year-old.
The half-time for movement of Pb from plasma to cortical bone surface is 14 minutes in an adult and
4 minutes in a 1-year-old. The major deposition fractions of Pb from plasma in the Leggett (1993) model
are to extravascular fluids (70%-74%) and RBCs (20%-24%). Slightly greater than the transfer from
plasma to soft tissues (8%-9%), which is minimally affected by age, the transfer from plasma to bone is
8% in adults and 14% for a 1-year-old (Leggett. 1993). Of the transfer from plasma to bone, trabecular
bone is expected to receive 56% of the Pb depositing in bone of adults and only 20% of the Pb depositing
in bone of 1-year-olds. Conversely, cortical bone receives 44% and 80% of Pb deposited in bone from
plasma in adults and 1-year-olds, respectively. Thus, the rates of transfer from plasma to bone and
compartmentalization between cortical and trabecular bone both vary with age.
When blood Pb concentrations are monitored in individuals over periods of years following a
cessation or decrease in exposure, the decrease in blood Pb concentration exhibits complex kinetics that
can be disaggregated into components that have faster and slower rates. The slower rates of clearance of
Pb from the blood over months and years following the cessation or reduction in exposures is thought to
primarily reflect elimination of Pb stores in bone. Nilsson et al. (1991) reported a tri-exponential decay in
the blood Pb concentrations of 14 individuals having a median occupational exposure period of 26 years.
Thirteen individuals had been temporarily removed from work because of excessive exposures (blood
levels >70 (ig/dL or high urinary S-aminolevulinic acid levels). Representing 22% of blood Pb, the fast
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compartment had a clearance half-time of 34 days. The intermediate compartment, 27% of blood Pb, had
a clearance half-time of 1.12 year. The slow compartment, 50% of blood Pb, had a clearance half-time of
13 years. The authors attributed the fast, intermediate, and slow compartment clearance to elimination of
Pb from blood and some soft tissues, from trabecular bone, and cortical bone, respectively. Rentschler et
al. (2012) also observed a slow terminal phase of Pb elimination from blood in five adults who had Pb
poisoning due to either occupational or nonoccupational exposures that ranged from approximately
1 month to 12 years and resulted in blood Pb concentrations of 70-110 (ig/dL. In this study, the blood Pb
monitoring period extended from 1 to 74 days following cessation of exposure to approximately 800 days
following the diagnosis of poisoning; however, it was not of sufficient duration to estimate the terminal
half-time. When the terminal half-time estimated by Nilsson et al. (1991) was used (13 years) to fit data
for these Pb poisoning cases to a two-component exponential decay model, the initial faster phase
represented approximately 80% of the blood Pb and the half-time was estimated to range from 60 to
120 days. The relatively longer fast phase half-time reported by Rentschler et al. (2012) compared with
Nilsson et al. (1991) may reflect the relatively high blood Pb concentrations in these poisoning cases that
resulted in temporary anemia and subsequent reestablishment of normal erythrocyte levels. In addition,
the use of a two-compartment model, with an assumed slow half-time of 13 years, as well as uncertainty
about the actual time of cessation of exposure may have prevented discerning a third, faster elimination
compartment in these data.
The longer half-life of Pb in bone compared with blood Pb, allows a more cumulative measure of
long-term Pb exposure. Pb in adult bone can serve to maintain BLLs long after external exposure has
ceased (Fleming et al.. 1997; Inskip et al.. 1996; Smith et al.. 1996; Kehoe. 1987; O'Flahertv et al.. 1982).
even for exposures that occurred during childhood (McNeill et al.. 2000). The more widespread use of in
vivo XRF Pb measurements in bone and indirect measurements of bone processes with stable Pb isotopes
have enhanced the use of bone Pb as a biomarker of Pb body burden.
Several studies have found a stronger relationship between patella Pb and blood Pb than tibia Pb
and blood Pb (Park et al.. 2009; Hu et al.. 1998; Hernandez-Avila et al.. 1996; Hu et al.. 1996). Hu et al.
(1998) suggest that trabecular bone is the predominant bone type providing Pb back into circulation under
steady-state and pathologic conditions. The stronger relationship between blood Pb and trabecular Pb
compared with cortical bone is probably associated with the larger surface area of trabecular bone
allowing for more Pb to bind via ion exchange mechanisms and more rapid turnover making it more
sensitive to changing patterns of exposure. Relationships between Pb in blood and bone in children and
adults are discussed in greater detail below (Sections 2.3.5.1, and 2.3.5.2).
2.3.5.1 Children
As discussed in Section 2.2.2.2, bone growth in children contributes to accumulation of Pb in
bone, which comprises most of the Pb body burden. As a result, bone Pb more closely reflects Pb body
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burden than blood Pb. However, changes in blood Pb concentration in children (i.e., associated with
changing exposures) are thought to more closely parallel changes in total body burden than such changes
in adults. Figure 2-6 shows a biokinetics model simulation of the temporal profile of Pb in blood and bone
in a child who experiences a period of constant Pb intake (from ages 2 to 5) via ingestion (fig Pb/day)
followed by an abrupt decline in intake. The figure illustrates several important general concepts about
the relationship between Pb in blood and bone. While blood Pb approaches a quasi-steady state after a
period of a few months with a constant rate of Pb intake (as demonstrated by the vertical dashed line), Pb
continues to accumulate in bone with continued Pb intake after the quasi-steady state is achieved in blood.
The model also predicts the rate of release of Pb from bone after a reduction in exposure is faster than in
adults. This difference has been attributed to accelerated growth-related bone mineral turnover in
children, which is the primary mechanism for release of Pb that has been incorporated into the bone
mineral matrix.
Several studies have examined blood Pb in children following changes in exposure. Children
(n = 3) removed from a relatively brief exposure to elevated environmental Pb exhibited faster slow-phase
kinetics than children (n = 3) removed from exposures that lasted several years, with half-times of 10 and
20-38 months, respectively (Manton et al.. 2000). The longer half-times measured under the latter
conditions reflect the contribution of bone Pb stores to blood Pb following a change in exposure.
However, the children exposed for the longer period (studied from ages 30 to 60 months) were older than
those exposed for a brief period (studied from ages 8 to 30 months), which may account for a portion of
the longer retention since bone remodeling decreases rapidly with age. Another study examined the time
for blood Pb to decrease below 10 (ig/dL in a large group of children (n = 579) having peak blood Pb at
an average age of 33 months (Roberts et al.. 2001). Children were grouped into four categories by their
peak blood Pb (10 to <15, 15 to <20, 20 to <25, and 25 to <30 (.ig/dL). The average time for the children's
blood Pb to decrease below 10 (ig/dL was 9.2, 14.3, 20.9, and 24 months, respectively. On the basis of the
mid-points of each blood Pb range and the time to reach 10 (ig/dL, the apparent half-time2 for clearance
from blood can be estimated at 29 months in the lowest blood Pb group and 16-18 months in the other
three groups. The increased half-life for the lowest blood Pb group suggests the lowest blood Pb group
may have experienced a longer duration of elevated Pb exposure than the three higher blood Pb groups.
A couple of studies investigated the relationship between blood and bone Pb in Pb-poisoned and
non-Pb-poisoned children recruited through Xinhua Hospital, Shanghai Jiaotong University, China
(Specht et al.. 2019b; Specht et al.. 2016). This discussion focuses on Specht et al. (2019b) and not their
preliminary results from fewer children (Specht et al.. 2016). K-XRF tibia bone Pb was well correlated
with blood Pb (r2 = 0.59; n = 157). The correlation between K-XRF and blood Pb improved (r2 = 0.95,
n = 24) by the time of a third chelation treatment. The authors attributed stronger correlation between
bone and blood Pb following the third chelation to a reduced effect in continued environmental Pb intake.
2Half-time (months) = [Ln(l/2)x(Time to 10 (ig/dL)] /Ln[(10 |ig/dL)/(midpoint blood Pb)|.
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Figure 2-7 illustrates the half-times for blood Pb in children reported in the study. Thirty-five percent of
the variability in half-times is attributable to the children's ages; sex was not influential. Children (9
females, 7 males) under the age of 3 had a fast half-time of only 6.4 days (SD: 3.5 days) that was
significantly (p < 0.001) less than observed in older children (8 females, 26 males; half-time: 19.2 days;
SD: 13.9).3 This study shows an equilibrium between Pb in bone and blood compartments in children
such that both are likely associated with total Pb body burden when continued environmental Pb intake
has been minimized or eliminated.
3These data were computed from supplemental data and differ slightly from values reported by the authors in their
paper. This difference appears to be due to the data of a 3.1 -year-old boy being grouped by the authors in data for
children <3 years of age.
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10
¦Blood
•Bone
Body
3.0
2 4 6
Age (year)
Note: Blood Pb concentration is thought to parallel body burden more closely in children than in adults, due to more rapid turnover of
bone and bone-Pb stores in children (upper panel). Baseline Pb intake is 3.2 |jg/day from birth until age 2, followed by a period of
increased intake (38.2 |jg/day) from age 2 until age 5, with a return to baseline intake of 3.2 |jg/day at age 5. The time-integrated
blood Pb concentration increases over time (lower panel). Simulation based on ICRP Pb biokinetics model CLeaaett. 19931 with
tissue and compartment masses and volumes based on equations and parameters from O'Flaherty's studies fO'Flahertv. 1995.
19931.
Figure 2-6 Simulation of relationship between blood Pb concentration and
body burden in children, with an elevated constant Pb intake from
age 2 to 5 years.
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Note: A quadradic fitted to the data is included to illustrate the trend in blood Pb half-times as a function of age. Linear and
quadradic functions both fit the data well (r2 = 0.35) and differ mainly in their intercepts of 1.7 and 3.9 days, respectively.
Source: Data for 50 children are from supplemental table ofSpecht etal. (2019b1.
Figure 2-7 Half-times of Pb in blood as reported by Specht et al. (2019b).
2.3.5.2 Adults
In adults, where a relatively large fraction of the body burden residing in bone has a slower
turnover compared with blood, a constant Pb uptake (or constant intake and fractional absorption) gives
rise to a quasi-steady state blood Pb concentration, whereas the body burden continues to increase over a
much longer period, largely because of continued accumulation of Pb in bone. This pattern is illustrated
by hypothetical simulations in Figure 2-8, wherein a low exposure to a constant baseline GI intake of
20 (ig/day occurs through the first 30 years of life. Subsequently, there is a 20-year period of increased
intake, wherein simulations show a relatively rapid increase in blood Pb concentration from a baseline of
approximately 2 (ig/dL to a new quasi-steady state, achieved in -75-100 days (i.e., approximately 3-4
times the blood elimination half-life). In contrast to the rapid increase in blood Pb, the bone and body
burden exhibit a steady increase across the full exposure 20-year period of enhanced exposure intake.
Following cessation of the 20-year enhanced exposure period at age 50, blood Pb concentration
declines rapidly compared with the slower decline in bone and body burden. There is a rapid drop in
blood Pb within a year from 9 to 3 (ig/dL (67% decrease) for the lower intake and from 90 to 40 (ig/dL
(55% decrease) for the higher intake exposure. Careful examination of the simulations shown in
Figure 2-8 reveals the accumulation and elimination phases of blood Pb kinetics are not symmetrical;
elimination is slower than accumulation as a result of the gradual release of bone Pb stores to blood. This
response, known as the prolonged terminal elimination phase of Pb from blood, has been observed in
retired Pb workers and in workers who continued to work after improved industrial hygiene standards
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reduced their exposures. These simulations in Figure 2-8 illustrate how a single blood Pb concentration
measurement or a series of measurements taken over a short time span could be a relatively poor index of
Pb body burden.
The drop in blood Pb concentrations following cessation of elevated exposure in Figure 2-8 is
well described (r = 0.996) by a tri-exponential decay function having the half-times of 30 days, 5 months,
and 8 years for the fast, intermediate, and slow compartments, respectively. For the low level of Pb intake
illustrated in the top panel of Figure 2-8, the fast, intermediate, and slow clearance compartments
represent 66%, 19%, and 15% of blood Pb, respectively. For the high level of Pb intake illustrated in the
bottom panel of Figure 2-8, the fast, intermediate, and slow clearance compartments represent 35%, 19%,
and 46% of blood Pb, respectively. The higher exposure resulted in more accumulation of Pb in bone
relative to the lower exposure scenario. This bone accumulation is reflected in the blood Pb clearance
kinetics by a larger slow compartment and smaller fast compartment for the high exposure.
One important potential implication of the profoundly different kinetics of Pb in blood and bone
is that, for a constant Pb exposure, Pb in bone will increase with increasing duration of exposure and,
therefore, with age. In contrast, blood Pb concentration will achieve a quasi-steady state. As a result, the
relationship between blood Pb and bone Pb will diverge with increasing exposure duration and age. This
divergence can impart different degrees of age-confounding when either blood Pb or bone Pb is used as
an internal dose metric in dose-response models. In a review of epidemiologic studies that evaluated the
associations between blood Pb, bone Pb, and cognitive function, the association was stronger for bone Pb
than blood Pb (particularly for longitudinal studies) for older individuals with environmental Pb
exposures and low BLLs (Shih et al.. 2007). In contrast, occupational workers with high current Pb
exposures had the strongest associations for BLLs with cognitive function, thus providing evidence for
this divergence (Shih et al.. 2007).
The expectation for an increase in bone Pb and body burden with age applies to scenarios of
constant exposure but not necessarily to real-world populations in which individual and population
exposures have changed over time. Longitudinal studies of blood and bone Pb trends have not always
found strong dependence on age (Nie et al.. 2009; Kim et al.. 1997). Kim et al. (1997) found bone Pb
levels increased with increasing age in elderly adults (age 52-83 years) only when the data were analyzed
cross-sectionally. When analyzed longitudinally, the trend for individual patella Pb was a 23% decrease
over a 3-year period (approximate ti/2 of 8 years), whereas tibia Pb levels did not change over the same
period. Therefore, although older individuals tended to have higher bone Pb levels, the 3-year temporal
trend for individuals was a loss of Pb from the more labile Pb stores in trabecular bone. Nie et al. (2011b)
observed longitudinal observations of blood and bone Pb in elderly adults did not show a significant age
effect on the association between blood Pb and bone Pb (patella and tibia), when the sample population
(n = 776) was stratified into age tertiles (mean age 62, 69 or 77 years).
Although differences in kinetics of blood and bone Pb degrade the predictive value of blood Pb as
a metric of Pb body burden, within a population that has similar exposure histories and age demographics,
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blood and bone Pb may show relatively strong associations. A recent analysis of a subset of data from the
Veterans Affairs (VA) Normative Aging Study (an all-male cohort) showed cross-sectional measurements
of blood Pb concentration accounted for approximately 9% (tibia) to 13% (patella) of the variability in
bone Pb levels. Inclusion of age in the regression model accounted for an additional 7%—10% of the
variability in bone Pb (Park et al.. 2009).
In addition to changes in exposure (discussed above), there are physiological processes in adults
during different life circumstances that can increase the contribution of bone Pb to blood Pb. These life
circumstances include times of physiological stress associated with enhanced bone remodeling, such as
during pregnancy and lactation (Hertz-Picciotto et al.. 2000; Silbergeld. 1991; Manton. 1985). menopause
or in the elderly (Silbergeld et al.. 1988). extended bed rest (Markowitz and Weinberger. 1990).
hyperparathyroidism (Kessler et al.. 1999) and severe weight loss (Riedt et al.. 2009).
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25 30 35 40 45 50 55 60 65 70
Age (year)
3
100 S.
25 30 35 40 45 50 55 60 65 70
Age (year)
Note: A constant baseline Gl intake of 20 |jg/day from age 0-30 results in a quasi-steady state blood Pb concentration and body
burden. An increase in Gl Pb intake to a relatively low intake of 120 |jg/day (top panel) or a high intake of 4,020 |jg/day (bottom
panel) from age 30 to 50 gives rise to a relatively rapid increase in blood Pb to a new quasi-steady state and a slower increase in
body burden. At age 50, intake returns to the baseline of 20 |jg/day. There is a rapid drop in blood Pb within a year from 9 to 3 |jg/dL
(67% decrease) for the lower intake and from 90 to 40 |jg/dL (55% decrease) for the high intake. As described in the text, the
decrease in blood Pb is well described by a tri-exponential decay function with higher intake having less in the fast compartment and
more in the slow compartment than the lower intake. Following the long period of high Pb intake, there is a rapid decline in blood Pb
over the first year followed by a more gradual decline in blood Pb. Simulation based on ICRP Pb biokinetics model CLeaaett. 19931
with tissue and compartment masses and volumes based on equations and parameters from O'Flaherty's studies fO'Flahertv. 1995.
19931.
Figure 2-8 Simulation of relationship between blood Pb concentration, bone
Pb, and body burden in adults.
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During pregnancy, bone Pb can serve as a Pb source as maternal bone is resorbed for the
production of the fetal skeleton (Gulson et al.. 2003; Gulsonetal.. 1999; Franklin et al.. 1997; Gulsonet
al.. 1997). Increased blood Pb during pregnancy has been demonstrated in numerous studies, and these
changes have been characterized as a "U-shaped" pattern of lower blood Pb concentrations during the
second trimester compared with the first and third trimesters (Lamadrid-Figucroa et al.. 2006; Gulson et
al.. 2004; Hertz-Picciotto et al.. 2000; Gulson et al.. 1997; Lagerkvist et al.. 1996; Schuhmacher et al..
1996; Rothenberg et al.. 1994). The U-shaped relationship reflects the relatively higher impact of
hemodilution in the second trimester versus the rate of bone Pb resorption accompanying Ca2+ releases for
establishing the fetal skeleton. In the third trimester, fetal skeletal growth on calcium demand is greater,
and Pb released from maternal skeleton offsets hemodilution. Gulson et al. (1998b) reported that during
pregnancy, blood Pb concentrations in the first immigrant Australian cohort (n = 15) increased by an
average of about 20% compared with nonpregnant migrant controls (n = 7). Skeletal contribution to blood
Pb, based on the isotopic composition of the immigrant subjects, increased in an approximately linear
manner during pregnancy. The mean increases for each woman during pregnancy varied from 26% to
99%. Interestingly, the percent change in blood Pb concentration was significantly greater during the post
pregnancy period than during the second and third trimesters. This is consistent with Hansen et al. (2011).
who demonstrated the greatest BLLs at 6 weeks postpartum compared with the second trimester in 211
Norwegian women. Increased calcium demands of lactation (relative to pregnancy) may contribute to the
greater change in blood Pb observed post pregnancy compared with the second and third trimesters. The
contribution of skeletal Pb to blood Pb during the post pregnancy period remained essentially constant at
the increased level of Pb mobilization.
Gulson et al. (2004) observed calcium supplementation was found to delay increased
mobilization of Pb from bone during pregnancy and halved the flux of Pb release from bone during late
pregnancy and postpartum. In another study, women whose daily Ca2+ intake was 850 mg per day showed
lower amounts of bone resorption during late pregnancy and postpartum than those whose intake was
560 mg per day (Manton et al.. 2003). Similarly, calcium supplementation (1,200 mg/day) in pregnant
Mexican women resulted in an 11% reduction in BLL compared with placebo and a 24% average
reduction for the most compliant women (Ettinger et al.. 2009). When considering baseline BLLs in
women who were more compliant in taking calcium supplementation, the reductions were similar for
those <5 (ig/dL and those >5 (ig/dL (14% and 17%, respectively). This result is in contrast to a study of
women who had blood Pb concentrations <5 (ig/dL, wherein calcium supplementation had no effect on
blood Pb concentrations (Gulson et al.. 2006). These investigators attributed their results to changes in
bone resorption with decoupling of trabecular and cortical bone sites.
Miranda et al. (2010) studied BLL among pregnant women aged 18-44 years old. The older age
segments in the study presumably had greater historic Pb exposures and associated stored Pb than the
younger age segments. Compared with the BLLs of a reference group in the 25- to 29-year-old age
category, pregnant women >30 years old had significant odds of having higher BLLs (ages 30-34:
OR = 2.39, p < 0.001; ages 35-39: OR = 2.98, p < 0.001; ages 40-44: OR = 7.69, p < 0.001). Similarly,
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younger women had less chance of having higher BLLs compared with the reference group (ages 18-19:
OR = 0.60, p = 0.179; ages 20-24: OR = 0.54, p = 0.015). These findings indicate maternal BLLs are
more likely the result of Pb mobilization of bone stores from historic exposures as opposed to
contemporaneous exposures.
BLLs increase during lactation due to alterations in the endogenous bone Pb release rate. After
adjusting for patella Pb concentration, an increase in BLLs of 12.7% (95% CI: 6.2, 19.6) was observed in
women who practiced partial lactation, and an increase of 18.6% (95% CI: 7.1, 31.4) was observed in
women who practiced exclusive lactation compared with those who stopped (Tellez-Roio et al.. 2002). In
another Mexico City study (Ettinger et al.. 2006; Ettinger et al.. 2004b). the authors concluded an
interquartile increase in patella Pb was associated with a 14% increase in breast milk Pb, whereas for tibia
Pb, the increase was -5%. Breast milk:maternal blood Pb concentration ratios are generally <0.1,
although values of 0.9 have been reported (Kovashiki et al.. 2010: Ettinger et al.. 2006: Gulson et al..
1998a). Dietary intake of polyunsaturated fatty acids (PUFA) has been shown to weaken the association
between Pb levels in patella and breast milk, perhaps indicating decreased transfer of Pb from bone to
breast milk with PUFA consumption (Arora et al.. 2008). Breast milk as a source of infant Pb exposure
was also discussed in Section 4.1.3.3 on dietary Pb exposure.
The Pb content in some bones (i.e., mid femur and pelvic bone) plateaus at middle age and then
decreases at older ages (Drasch et al.. 1987). This decrease is most pronounced in women and may be due
to osteoporosis and release of Pb from resorbed bone to blood (Gulson et al.. 2002). Two studies indicate
the endogenous release rate in postmenopausal women ranges from 0.13 to 0.14 (ig/dL in blood per |ig/g
bone and is nearly double the rate found in premenopausal women (0.07-0.08 (ig/dL per |ig/g bone)
(Popovic et al.. 2005: Garrido Latorre et al.. 2003). An analysis of data on blood Pb concentrations and
markers of bone formation (serum alkaline phosphatase) and resorption (urinary cross-linked
N-telopeptides, NTx) in a sample of U.S. women found that blood Pb concentrations were higher in
women (pre- or postmenopausal) who exhibited the highest bone formation or resorption activities
(Jackson et al.. 2010). Calcium or vitamin D supplementation decreased the blood Pb concentrations in
the highest bone formation and resorption tertiles of the population of postmenopausal women.
Significant associations between increasing NTx and increasing BLLs (i.e., increased intercept of
regression model relating the change in blood Pb per change in bone Pb) have also been observed in
elderly men (Nie et al.. 2009).
Studies of the effect of hormone replacement therapy on bone Pb mobilization have yielded
conflicting results (Popovic et al.. 2005: Berkowitz et al.. 2004: Garrido Latorre et al.. 2003: Korrick et
al.. 2002: Webber et al.. 1995). In women with severe weight loss (28% of body mass index [BMI] in
6 months) sufficient to increase bone turnover, increased BLLs of approximately 2.1 (ig/dL (250%) were
reported, and these blood Pb increases were associated with biomarkers of increased bone turnover
(e.g., urinary pyridinoline cross-links) (Riedt et al.. 2009).
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2.3.6
Relationship between Pb in Blood and Pb in Soft Tissues
Figure 2-9 shows simulations of blood and soft tissue Pb (including brain) for the same exposure
scenarios previously displayed. Pb uptake and elimination in soft tissues is much faster than in bone. As a
result, following cessation of a period of elevated exposure, Pb in soft tissues is more quickly returned to
blood. The terminal elimination phase from soft tissue mimics that of blood, and it is similarly influenced
by the contribution of bone Pb returned to blood and being redistributed to soft tissue.
Information on Pb levels in human brain is limited to autopsy data. These data indicate
brain/blood Pb ratios of approximately 0.5 in infancy, which remain relatively constant over the lifetime
(range 0.3 to 1.1) (Barry. 1981; Barry. 1975). The simulation of brain Pb shown in Figure 2-10 reflects
general concepts derived from observations made in nonhuman primates, dogs, and rodents. These
observations suggest peak Pb levels in the brain are reached 6 months following a bolus exposure, and
within 2 months, approximately 80% of steady state brain Pb levels are reached (Leggett. 1993). There is
a relatively slow elimination of Pb from brain (ti/2 ~ 2 years) compared with other soft tissues (Leggett.
1993). This slow elimination rate is reflected in the slower elimination phase kinetics shown in
Figure 2-10. Although in this model, brain Pb to blood Pb transfer half-times are assumed to be the same
in children and adults, uptake kinetics are assumed to be faster during infancy and childhood, which
achieves a higher fraction of the soft tissue burden in brain, consistent with higher brain/body mass
relationships. The uptake half-times predicted by Leggett (1993) vary from 0.9 to 3.7 days, depending on
age. Brain Pb kinetics represented in the simulations are simple outcomes of modeling assumptions and
cannot currently be verified with available observations in humans.
Urinary filtering and excretion of Pb is associated with plasma Pb concentrations. Given the
curvilinear relationship between blood Pb and plasma Pb, a secondary expectation is for a curvilinear
relationship between blood Pb and urinary Pb excretion that may become evident only at relatively high
blood Pb concentrations (e.g., >25 (.ig/dL). Figure 2-11 shows these relationships predicted from the
model. In this case, the exposure scenario shown is for an adult (age 40 years) at a quasi-steady state PbB;
the same relationships hold for children (Leggett. 1993). At lower blood Pb concentrations (<25 (.ig/dL).
urinary Pb excretion is predicted to closely parallel plasma Pb concentration for any given BLL
(Figure 2-11, top panel). It follows from this that, similar to blood Pb, urinary Pb will respond much more
rapidly to an abrupt change in Pb exposure than will bone Pb. One important implication of this
relationship is, as described previously for blood Pb, the relationships between urinary Pb and bone Pb
will diverge with increasing exposure duration and age, even if exposure remains constant. Furthermore,
following an abrupt cessation of exposure, urine Pb will quickly decrease while bone Pb will remain
elevated (Figure 2-11, lower panel).
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Note: For the child simulation (upper panel), baseline Pb intake is 3.2 |jg/day from birth until age 2, followed by a period of increased
intake to 38.2 |jg/day from age 2 to 5, with a return to baseline intake at age 5. For the adult simulation (lower panel), baseline
intake is 20 |jg/day from age 0 to 30, followed by a 20-year period of increased intake to 120 |jg/day from age 30 to 50, with a return
to baseline intake at age 50. Simulation based on ICRP Pb biokinetics model CLegqett. 19931 with tissue and compartment masses
and volumes based on equations and parameters from O'Flaherty's studies fO'Flahertv. 1995. 19931.
Figure 2-10 Simulation of blood and brain Pb in children and adults who
experience a period of increased Pb intake.
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20 30
Blood Pb (|ig/dL)
12
10 ¦
>_i
ra -a
So 8
ra 3.
3.—
Q- 6
-------
2.4
Studies of Pb Biomarker Levels
This section provides information on studies containing Pb biomarker concentrations, including
in blood, bone, urine, teeth, and others. NHANES data show BLLs have continued on a downward trend
since 1976. EBLLs have been linked in the literature to air sources (including proximity to airports), soil
and dust (including from housing demolition and older homes), dietary sources, and tap water such as in
the case of the Flint Water Crisis, among other sources described in 2.1. Continued research since U.S.
EPA (2013) has shown there is a seasonality component to BLLs linked to several factors, including
higher resuspension rates of soil containing Pb during drier months.
2.4.1 Pb in Blood
As concluded in the 2013 Pb ISA (U.S. EPA. 2013). trends in BLLs have been decreasing for
U.S. residents over the past 45 years, as evidenced by NHANES data. Data show a progressive downward
trend has occurred during the 1976-2018 period. The 2013 Pb ISA (U.S. EPA. 2013) noted the most
dramatic declines occurred coincident with the phase-out of leaded gasoline and reductions in point
source Pb emissions. The temporal trend for GM BLLs by age group from the 1999-2018 period is
shown below in Figure 2-12. Summary statistics from the National Report on Human Exposure (CDC.
2021b) containing NHANES BLLs from 2011 to 2018 is presented in Table 2-11 below. In agreement
with study results presented in Section 2.1.5.4, Figure 2-14 shows the gap in BLLs between non-Hispanic
Black children and children of different racial/ethnic groups has decreased over time.
Table 2-11
Blood-Pb concentrations in the U.S. population
Survey Stratum
Period
Geometric Mean
(Hg/dL)
95% Confidence
Interval
Number of
Subjects
All
2011-2012
0.973
0.916, 1.04
7,920
2013-2014
0.858
0.813, 0.906
5,215
2015-2016
0.820
0.772, 0.872
4,988
2017-2018
0.753
0.723, 0.784
7,513
1-5 yr
2011-2012
0.970
0.877, 1.07
713
2013-2014
0.782
0.705, 0.869
818
2015-2016
0.758
0.675, 0.850
790
2017-2018
0.670
0.600, 0.748
629
6-11 yr
2011-2012
0.681
0.623, 0.744
1,048
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Survey Stratum
Period
Geometric Mean
(Mg/dL)
95% Confidence
Interval
Number of
Subjects
2013-2014
0.567
0.529, 0.607
1,075
2015-2016
0.571
0.523, 0.623
1,023
2017-2018
0.475
0.456, 0.494
833
12-19 yr
2011-2012
0.554
0.511, 0.601
1,129
2013-2014
0.506
0.464, 0.551
627
2015-2016
0.467
0.433, 0.504
565
2017-2018
0.411
0.387, 0.436
1,030
>20 yr
2011-2012
1.090
1.03, 1.16
5,030
2013-2014
0.967
0.921, 1.02
2,695
2015-2016
0.920
0.862, 0.982
2,610
2017-2018
0.855
0.816, 0.895
5,021
Male
2011-2012
1.130
1.06, 1.21
3,968
2013-2014
0.994
0.919, 1.08
2,587
2015-2016
0.921
0.864, 0.981
2,488
2017-2018
0.860
0.820, 0.902
3,666
Female
2011-2012
0.842
0.796, 0.890
3,952
2013-2014
0.746
0.715, 0.777
2,628
2015-2016
0.735
0.679, 0.795
2,500
2017-2018
0.664
0.632, 0.698
3,847
Mexican American People
2011-2012
0.838
0.767, 0.916
1,077
2013-2014
0.746
0.685, 0.813
969
2015-2016
0.704
0.659, 0.752
994
2017-2018
0.662
0.610, 0.719
1,134
Non-Hispanic Black People
2011-2012
0.998
0.947, 1.05
2,195
2013-2014
0.871
0.787, 0.963
1,119
2015-2016
0.856
0.763, 0.962
1,070
2017-2018
0.766
0.736, 0.798
1,708
Non-Hispanic White People
2011-2012
0.993
0.914, 1.08
2,493
2013-2014
0.882
0.820, 0.950
1,848
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Survey Stratum
Period
Geometric Mean
(Hg/dL)
95% Confidence
Interval
Number of
Subjects
2015-2016
0.835
0.774, 0.900
1,511
2017-2018
0.772
0.731, 0.816
2,536
All Hispanic People
2011-2012
0.855
0.793, 0.922
1,931
2013-2014
0.742
0.695, 0.793
1,481
2015-2016
0.703
0.658, 0.750
1,664
2017-2018
0.629
0.593, 0.667
1,816
Asian People
2011-2012
1.150
1.06, 1.24
1,005
2013-2014
1.010
0.923, 1.11
510
2015-2016
1.070
0.976, 1.18
479
2017-2018
1.020
0.909, 1.15
946
yr = year(s).
Limits of detection (LOD) for survey years 11-12, 13-14, 15-16, and 17-18 are 0.25, 0.07, 0.07, and 0.07, respectively.
Source: Data sourced from CDC ('2021a').
2.75
2.25
T3
J! 175
TS
(0
• < '
* i *
• Children 1-5 yrs
A Children 6-11 yrs
~ Teens 12-19 yrs
¦ Adults >20 yrs
99-00 01-02 03-04 05-06 07-08 09-10 11-12 13-14 15-16 17-18
Survey Period
Note: Shown are geometric means and 95% Cis based on data from NHANES IV CDC ('2021a1.
Figure 2-12 Temporal trend in blood Pb concentrations.
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COHORT BIRTH YEARS
— 1900-1929
— 1930-1939
— 1940-1949
— 1950-1959
1960-1969
— 1970-1974
— 1975-1979
— 1980-1984
— 1985-1989
1990-1994
— 1995-1999
— 2000-2004
— 2005-2009
— 2010-2014
EXAM YEAR
Note: The means of logged blood Pb were weighted to represent national averages. Data were from the publicly available NHANES
](, NHANES III, and continuous NHANES cycles (1999-2000, 2003-2004, 2005-2006, 2007-2008, 2011-2012, 2013-2014, 2015-
2016, 2017-2018). 2001-2002 and 2009-2010 were excluded because only 551 blood Pb samples were available for each,
respectively. Data from 2015 to 2016 for birth cohort 2015-2016 was excluded from the figure due to small sample size (n = 49;
participants with available blood Pb data).
Figure 2-13 Blood Pb cohort means versus year of exam.
3.0
l,2-5
a 2.0
0)
m
£ 1.5
¦4—"
0
1.0 H
„cA „cA .ofiP„«
Survey Year
Survey Year
Race/Ethnicity
• Mexican American
• Other Hispanic
• Non-Hispanic White
• Non-Hispanic Black
Other Race - including Multi-Racial
Figure 2-14 Blood Pb geometric means versus year of NHANES exam by
race/ethnicity.
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Additional analyses have used NHANES data to investigate the decline in BLLs over time. Wang
et al. (2021) analyzed NHANES data from 1996 to 2016 that included 68,877 participants (1-85 years;
38-year weighted mean age) and found an annual percentage change of-4.26% (p < 0.05) during this
time period from a mean BLL of 1.68 (ig/dL (95% CI: 1.63, 1.74) to 0.82 (ig/dL (95% CI: 0.77, 0.87).
Ettinger et al. (2020) analyzed BLLs of women of childbearing age (15-49 years) using 1976-2016
NHANES data (n = 22,408). The authors found the GM of this group dropped over a 40-year period from
10.37 (ig/dL (95% CI: 9.95, 10.79) to 0.61 ^ig/dL (95% CI: 0.59,0.64) (from 1976-1980 to 2011-2016,
respectively). Few women (0.7%) in the 2011-2016 group had BLLs above 5 (ig/dL. By comparison, for
children aged 1-5 years, the 1976-1980 NHANES showed ablood Pb of 15.2 (ig/dL (95% CI: 14.3, 16.1)
with nearly all (99.8%) exceeding 5 (ig/dL, which declined in 2011-2016 to 0.8 (ig/dL (95% CI: 0.8, 0.9)
with only 1.3% exceeding 5 (ig/dL (Egan et al., 2021).
The 1986 and 2006 Pb AQCDs (U.S. EPA. 2006. 1986) and the 2013 Pb ISA (U.S. EPA. 2013)
contain evidence that BLLs may follow a seasonal pattern in children, with elevated concentrations in the
warm season compared with lower levels in the cold season. This is important to understand when studies
reporting BLLs are evaluated because seasonal effects may also contribute to findings, especially at low
BLLs, wherein contributions from other sources may have a greater impact on BLLs than the source
being studied, potentially serving as a confounding variable when the link between BLLs and an exposure
pathway is being investigated.
Levin et al. (2020) reviewed literature within the previous Pb ISA and AQCD documents for
information on Pb seasonality and supplemented conclusions based on recent research. The authors found
seasonality of BLLs could be linked to multiple sources including gasoline usage (prior to 1985), soil and
dust, housing renovations, avgas, drinking water, diet, consumption of game meat, off-road vehicles, and
vitamin D generation. As mentioned in Section 1.2.6 of this document
(https://assessments.epa.gov/isa/document/&deid=359536), there is evidence that soil resuspension
occurs to the greatest degree during the summer and fall when there are drier soil conditions (Resongles et
al.. 2021; Mielke et al.. 2019b; Laidlaw et al.. 2017b; Laidlaw et al.. 2016; Laidlaw et al.. 2014). Laidlaw
et al. (2012) concluded soil contributions to atmospheric Pb were highest during the summer and fall in
Pittsburgh, PA; Detroit, MI; Chicago, IL; and Birmingham, AL. Laidlaw et al. (2016) investigated
seasonality in child BLLs in Flint, MI, finding children's average BLLs consistently peaked in the third
quarter of the year between 2010-2015, and concluded there was likely a contribution of soil Pb through
resuspension to child BLLs. Zahran et al. (2013a) investigated soil contributions of Pb to atmospheric
levels and the effect of this atmospheric Pb on 367,839 BLLs in children in Detroit. Atmospheric soil was
derived using a mineral equation based on the elemental composition of soil (Al, Si, Ca, Fe, and Ti), and a
regression model was built between atmospheric soil and atmospheric Pb concentrations that included an
adjustment using local weather conditions (including humidity, sea level pressure, temperature, visibility,
and wind speed). After controlling for child sex, blood draw type, and year of observation, the authors
found an increase of one standard deviation in air Pb (-0.0006 (ig/m3) was associated with an 8.04%
(95% CI: 7.1 to 9.0%) increase in BLL for children less than 1 year of age. In addition, it was found that
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after adjusting for local weather conditions, one of the models showed an air Pb increase of 0.39% (95%
CI: 0.28%, 0.50%) for every 1% increase in atmospheric soil (i.e., resuspended soil). Atmospheric soil
was found to be a stronger contributor to air Pb than road dust. The study also found that, with the
exception of children aged 3, absent soil resuspension, air Pb had little observable effect on child BLLs.
This condition was assessed by regressing child BLL on the residual of their model, which represents
other unmeasured sources of air Pb present.
Shao et al. (2017) analyzed 83,127 BLL data records of children in Syracuse, NY collected by the
Onondaga County Health Department from April 1992 to December 2011. The authors found
interventions by the Syracuse Lead Program to remove Pb-based paint in the homes of those children with
BLLs over 10 (ig/dL resulted in a change of the seasonal peak of BLLs from the summer (June, July,
August) to different months, without a consistent pattern by year. This suggested BLL seasonality may
have been influenced more by Pb-based paint exposure through opening of windows for natural
ventilation in the summer (while being left closed in the winter) than by other factors, such as time
outdoors and increased exposure to Pb in soil or dust from soil. Past studies found a regular seasonal peak
in summer for BLLs of Syracuse children (Laidlaw et al.. 2005; Haley and Talbot 2004).
The Cochrane Library includes several systematic reviews and meta-analyses of randomized
controlled trials and quasi-randomized controlled trials (Nussbaumer-Streit et al.. 2020; Nussbaumer-
Streit et al.. 2016; Yeoh et al.. 2014; Yeoh et al.. 2008). The studies included in these reviews were
conducted to evaluate the effectiveness of interventions, including dust control actions like soil
excavation and replacement that were intended to reduce children's BLLs. Overall, these reviews found
no statistical evidence that these interventions were effective in reducing children's BLL. The authors
also noted the evidence specifically pertaining to the effect of soil remediation was limited to two studies
(Farrell et al.. 1998; Weitzman et al.. 1993) reporting contradictory findings.
Braun et al. (2018) conducted an intervention study related to indoor dust Pb in Cincinnati, OH
between 2003 and 2006. Pregnant women were randomly assigned to a residential Pb hazard intervention
group (n = 174) or a control group (n = 181). The former received interventions such as the covering of
bare yard soil areas, repair of deteriorated Pb-based paint areas, and extensive dust control and cleanup,
whereas the latter received injury prevention education. Both the control and intervention groups showed
reductions in floor, windowsill, and window trough dust Pb loading during the study. The BLLs of
children from 1 to 8 years of age were not significantly different between the control and intervention
groups, although the geometric mean childhood blood Pb levels were higher in non-Hispanic Black
children (Figure 2-15). However, the reductions in dust Pb loadings of the control group during the 2-year
period following inclusion in the study suggests that the control group may have been influenced by
participation in the study so as to undertake measures to reduce dust Pb loadings within their residences.
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I A All children
4.10-j
3.00-
2.50-
2.00-
-o
s
:
1.50-
1.00-
0.36-
3 5 S
Child Age, y
10
BI Non-Hispanic black
4.10-
0.361
1 3 S 8
Child Age, y
10
c ] Non-Hlspanfc white
4-IOt
2.50-
| Intervention group
, Control group
13 5 8 10
Child Age, y
Note; Age-specific geometric mean blood Pb levels were derived from a mixed model that included the intervention arm, a 5-knot
cubic polynomial spline for age and intervention by age interactions. Shading indicates 95% confidence intervals. BLLs are reported
in pg/dL. Source: Braun et al. (2018)
Figure 2-15 Geometric mean childhood blood Pb levels assessed between 1
and 8 years old, stratified by race/ethnicity.
Ye et al. (2022) assessed the effect of soil remediation on BLLs of children living in Omaha, ME.
Children's BLLs within a 27 mi2 study area were paired with residential yard soil Pb and remediation
status. A 13 mi2 focus area within the study area delineated where soil Pb concentrations for at least 1 in
20 homes exceeded 400 ppm. Blood Pb data were available for nearly 75,000 children (0-7 years old)
living within the study area between 1999 and 2016. Residential soil Pb data were available for 14,000
non-remediated properties and 7,400 properties that received remediation. Before remediation, children's
risk of having an EBLL (i. e., >5 ug/dL) was associated with both residential soil Pb [OR = 2.00; 95%
confidence interval (CI): 1.83, 2.19; >400-800 versus <200 ppm] and neighborhood soil Pb [OR= 1.85
(95% CI: 1.62, 2.11; >400-800 versus <200 ppm)]. The odds of having an EBLL was higher before
remediation than after [OR = 1.52 (95%CI: 1.34, 1.72)]. This study showed a benefit of soil Pb
remediation in reducing the risk of EBLLs in children, but the effects of activities such as community
surveillance and health education may have contributed, in part, to this benefit.
As mentioned in Section 1.2 of this document ( tps://assessments.epa.gov/isa/document/
&deid=359536). aviation fuel remains a major source of Pb emissions in ambient air. M iranda et al
(20111 found a monotonically decreasing trend in BLLs of children aged 9 months to 7 years and distance
from airports in six counties in North Carolina, although emissions were not included in the model.
Children within 500 m, 1,000 m, and 1,500 m had BLLs that were on average 4.4%, 3.8%, and 2.1%
higher, respectively, than other children in those counties. Zahran et al. (2017a) analyzed the BLLs of
1,043,391 children, aged 1-5 years, collected from January 2001 through December 2009. Blood samples
were collected during doctors" visits with a sampling emphasis on at-risk children in older homes or
neighborhoods with EBLLs. BLLs were linked spatially and temporally to 448 airports in Michigan and
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emission inventories of Pb-releasing industry sites found in the TRI. Measurements at/below detection
levels were found to be 40.2%. The authors found a 3.4% reduction in the odds of surpassing a CDC
threshold for EBLL (>5 and 10 (ig/dL) with each km distance of a child's residence from an airport. The
authors also found an increased likelihood of exceeding a 5 (ig/dL BLL to be associated with increases of
100 piston engine aircraft operations per month; this effect decreased about 1% for every 1-km increase in
residence distance from the airport. In another study, BLLs were higher for Republic of Korea Air Force
crews working at bases using avgas (4.20 (ig/dL) compared with those using jet propellant (3.79 (.ig/dL).
with correlations also observed between BLL and longer working hours on airport runways at the bases
using avgas (Park et al.. 2013).
Zahran et al. (2023) analyzed 14,000 blood Pb concentration samples for children <5 years of age
in neighborhoods surrounding Reid-Hillview Airport, Santa Clara County, CA during a 10-yr observation
period from January 2011-December 2020. In addition, three potential indicators of avgas exposure were
investigated, including (1) child residential distance from the airport, (2) whether the child's residence
was downwind, and (3) volume of piston-engine aircraft traffic from the date of the blood draw. The
authors found that the odds of a child's BLL exceeding 4.5 (ig/dL increased statistically significantly (p
<0.05) with proximity to the airport, for those living east and predominantly downwind, and with
increasing volume of piston-engine aircraft traffic. Model results were controlled for the number of
U.S. EPA TRI facilities < 2 miles of a child's residence, use of Pb-based paint in homes as indicated by
the percentage of homes built in the neighborhood before 1960 (i.e. when use of Pb-based paint had
declined by more than 90% from peak usage in the 1920s) at the year of blood draw, and SES as indicated
by percentage of adults with a college degree, median home prices, and median household incomes.
Hollingsworth and Rudik (2021) analyzed ambient air Pb concentrations and EBLLs in relation to
leaded gasoline usage in automotive races. The percentages of BLLs above 10 (ig/dL in children under
72 months old for a given county-year were retrieved from CDC State Surveillance data, for which blood
Pb sampling was targeted in high-risk areas. Only confirmed cases of EBLLs were used, either a venous
blood draw showing a BLL above 10 (ig/dL or two capillary draws within two weeks of each other
showing a BLL above 10 (ig/dL. Using an event study and spatial lag model, the authors estimated that
every 100,000 miles driven using leaded gasoline in the previous week resulted in an increase in mean
concentrations of ambient Pb within 50 miles of a racetrack equivalent to 10 percent. Data within Figure 4
of the paper shows that the average prevalence of EBLLs for border counties (i.e., counties bordering
those where races occurred) and race counties (i.e., counties where races occurred) were one and two
percentage points higher than control counties before 2007, respectively, when the National Association
for Stock Car Auto Racing and the Automobile Racing Club of America switched to using unleaded fuel.
After 2007, the prevalence of EBLLs was similar to control counties. In addition, the authors developed a
regression model linking the prevalence of EBLLs to whether there was a race in a county and whether a
county bordered another county where a race occurred. This model also included a set of controls for SES
(as indicated by unemployment rate, median income, percent non-white in the county), payroll in the
manufacturing sector, and quantity of TRI Pb emissions. They estimated that the effects of living in a race
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county on EBLL prevalence were higher by 18 and 13 percent, in 2005 and 2006 respectively, than 2007.
This drop from 2005 to 2006 was consistent with the fact that 14 percent of race miles driven in 2006
used unleaded gasoline. These results suggest that leaded gasoline usage in automotive races prior to 2007
led to a greater prevalence of EBLLs in counties that had races as well as bordering counties.
Meng et al. (2014) used 1999-2008 NHANES BLL data and merged it with contemporaneous Pb
air concentrations from the U.S. EPA AQS at monitors within 4 km of NHANES participants. The
authors generally found positive associations between BLLs for all five age groups (1-5, 6-11, 12-19,
20-59, and >60 years) and Pb concentrations at Pb-PMio monitors. This study is described in greater
detail in Section 2.5.
Soil and dust have been investigated for contributions to BLLs. As mentioned in Section 2.1.3.2,
Mielke et al. and other research teams have published a series of papers (Mielke et al.. 2019b; Mielke et
al.. 2019a. 2017; Rabito et al.. 2012; Mielke et al.. 2011b; Zahran et al.. 2011) demonstrating the
importance of soil Pb as a source of children's Pb exposures in New Orleans and other cities. The New
Orleans data they developed was especially extensive (>5,000 surface soil samples; >50,000 blood Pb
samples) and have included multiple time points demonstrating a now declining pattern of soil Pb
concentrations and BLLs. Correlations were found between soil Pb levels and BLLs in children both
before and after Hurricane Katrina.
Stewart et al. (2014) used 81 soil samples and bioavailability data to predict BLLs using the
IEUBK model in Toledo, OH. EBLLs for the 1-2-year-old age group were predicted in 28.4% of areas
sampled. Pavilonis et al. (2022) collected 1,504 soil samples from 43 parks in Brooklyn, NY and EBLL
information on children aged 1-5 years made available by the New York City DOHMH. The rate of
EBLLs per 1,000 children was highest in the locations within the highest quartile of soil Pb
concentrations (>150 ppm, mean rate: 42.4, median rate: 37.2). The authors did not see a monotonic
increase in the rate of EBLL by quartile; however, a multivariable regression model that controlled for
race/ethnicity and housing characteristics found a significant positive association between soil Pb
concentrations and EBLL rates (p = 0.004). Morrison et al. (2013) collected 226 soil samples around
neighborhoods in Marion County, IN. The authors analyzed these in relation to 16,232 BLL records of
children living within the county. The authors found no statistical association between soil Pb
concentrations and BLLs at the census block level; however, children within the urban core of the county
were more likely to have EBLLs, likely due to traffic and industrial sources.
Bradham et al. (2017) analyzed the relationship between total Pb soil concentration, bioaccessible
Pb soil concentration (38 soil samples), and BLLs of children aged 1 to 7 years (49 children) around
Philadelphia residential homes. Regression models developed by the authors found the use of total soil Pb
as a predictor of BLL variability accounted for 23% of variability, whereas the use of bioaccessible soil
Pb accounted for 26% of variability (R2 value 0.23 versus 0.26), suggesting bioaccessible soil Pb
concentrations may be a better predictor of child BLLs.
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To improve preventive methods for Pb exposure from soil and dust Zahran et al. (2013b)
performed a study looking at the importance of soil sample locations in predicting child Pb exposure. Soil
samples (n = 5,467) were collected across 286 census tracts and compared against geo-referenced blood
Pb data of 55,551 children in New Orleans. The authors found the strongest soil type predictor of
between-neighborhood variation in BLLs was residential street soils (39.7%), followed by busy street
soils (21.97%), open space soils (20.25%), and home foundation soils (18.71%). The authors concluded
the turbulent environment created by roadways leads to resuspension of dust in soils, increasing
accidental inhalation and ingestion of Pb in those soils.
Several studies have investigated the link between housing, demolitions, and BLLs. Eisenberg et
al. (2020) found children younger than 6 years old living in Detroit, MI, in homes previously foreclosed
on (which tend to be older and may be less likely to receive Pb remediation actions) were more likely to
have EBLLs. Eighty-four percent of the sample population lived in housing built before 1950. Ninety-
three percent of children having EBLLs lived in older housing. Thirteen percent of children living in
housing near two or more recent demolitions had EBLLs compared with <8% of children who did not live
near recent demolitions. Clark et al. (2011) found BLLs declined up to 3 years after housing interventions
were enacted to control Pb-based paint standards in low-income, privately owned housing. Chiofalo et al.
(2019) compared the BLLs of 4,693 children in New York City and found 2.76% of children in private
housing had EBLLs while 0.25% in public housing had EBLLs. Most of the public housing was built
before 1960; however, ZIP codes for private housing with the most children who had BLLs at or above
5 (ig/dL had a high prevalence of older housing as well. McClure et al. (2016) analyzed 5,266,408 BLLs
of children <6 years of age from May 2009 to April 2015 in 36 states. The authors found that living in
ZIP codes with >51.0% of homes built before 1950 had a significantly larger association with BLLs
>5.0 ng/dL (OR 5.86, 95% CI: 5.71-6.01) or >10 ^ig/dL (OR 6.34, 95% CI: 5.97-6.74) than living in ZIP
codes with <3.6% of homes built before 1950.
Bezold et al. (2020) investigated the association of BLLs in children <6 years of age (n = 54,150
BLL observations) with demolition activities within 400 feet of their homes during an uptick of
demolitions within Detroit, MI in 2014-2018. The authors found associations between EBLLs (>5 (ig/dL)
and housing demolitions for the years between 2014 and 2017 but not 2018 (p = 0.07), which the authors
attributed to differences in dust management practices between years, and the fact that homes demolished
in 2018 were, on average, newer than those demolished earlier. Spanier et al. (2013) surveyed parents of
276 children in the Rochester area about renovation activities and related it to children's BLLs. It was
found that interior housing renovation activities were associated with a 12% increase in children's BLLs.
Dignam et al. (2019) performed a study to identify risk factors associated with EBLLs among 104
children in Philadelphia neighborhoods. Higher GM BLLs were significantly associated with door Pb
content >40 (ig/ft2 (p = 0.0027) and living in a home built before 1980 (p = 0.0017).
As discussed in the 2013 Pb ISA (U.S. EPA. 2013). consumption of Pb-contaminated material,
including soil, paint, drinking water, and food, has been linked to increased BLLs. A study of 491
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pregnant women in New York City found those who reported engaging in pica had BLLs on average
higher than those who did not report pica behavior (29.5 versus 23.8 (ig/dL, p < 0.0001) and those
engaging in pica were 11 times more likely to receive chelation therapy (Thihalolipavan et al.. 2013).
Keller et al. (2017) investigated factors that contributed to BLLs over 45 (ig/dL in 145 children in New
York City during the period between 2004 and 2010. The strongest reported risk factor was eating paint
(36%), followed by other risk factors such as spending time outside the United States (34%) and having a
developmental delay (27%). Children with developmental disorders may behave like younger children,
with hand-to-mouth activity that persists longer, leading to greater exposure through ingestion of Pb over
time (Shannon and Graef. 1996).
Desai et al. (2021) used 2009-2014 NHANES data for 12- to 36-month-olds to investigate the
existence of a link between foods consumed and BLLs. They found that while consumption of the
majority of food groups showed little effect on BLLs, cereal and milk consumption was associated with
lower BLLs, whereas meat and fruit juice consumption was linked to higher BLLs. Wang et al. (2017a)
found higher intakes of processed meat, red meat, refined grains, high-fat dairy products, French fries,
butter, and eggs were associated with higher levels of BLLs in middle-aged to elderly men, using data
from the VA Normative Aging Study. Savadatti et al. (2019) found that among a sample of licensed
anglers and Burmese immigrants in Buffalo, NY, those who were more likely to catch and consume local
fish had higher GM BLLs than 2013-2014 reference levels. Davis et al. (2014) examined the associations
between 49 foods and biomarkers of Pb, Hg, Cd, and As in NHANES participants. They found diet
explained a 2.9% variation in blood Pb in children and a 1.6% variation in adult BLLs. The authors
acknowledged dietary data were self-reported, meaning participants may have been subject to
misclassification, and their results cannot be generalized to the U.S. population. Colapinto et al. (2016)
investigated tea consumption in 1,954 pregnant Canadian women and found increased tea consumption
was linked to higher BLLs. However, the GMs of women who consumed the greatest amount of tea were
less than 1 (ig/dL.
Pb found in drinking water during the Flint Water Crisis has been associated with EBLLs in
children. During the period April 25, 2014-October 15, 2015, the water source for residents in Flint, MI
was switched from Lake Huron to the Flint River (Kennedy et al.. 2016). Hanna-Attisha et al. (2016)
examined BLLs in children <5 years of age before and during the rise in tap water Pb concentrations
(n = 1,473; pre = 736; post = 737). They found the incidence of BLLs at or above 5 (ig/dL increased from
2.4% to 4.9% from 2013 to mid-2015. Neighborhoods with the highest Pb concentrations in tap water
experienced a 6.6% increase in EBLLs. In contrast, for neighborhoods outside the city that did not receive
water treated at the Flint facility, there was no statistically significant (p < 0.05) change in incidence of
EBLLs. Alternative potential Pb exposure sources, such as demolition projects, new Pb-producing
factories, changes in Pb remediation programs, or manufacturing that uses Pb, showed no spatial
relationship to increased BLLs. Gomez et al. (2018) found that among 15,817 BLLs for children <5 years
of age, the GM decreased from 2.33 (ig/dL in 2006 to 1.15 (ig/dL in 2016; however, during that decade,
the GM increased twice, once in 2010-2011 (a period before the switch to Flint River water) and again in
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2014-2015 (during the Flint Water Crisis). By analyzing BLLs of children <6 years old from April 2013
to March 2016, Kennedy et al. (2016) found that by analyzing BLLs of children less than six years old
from April 2013-March 2016 that 3.0% of BLLs were above 5 (ig/dL. The percentage of children with
EBLLs in the period of April 2014-January 2015, before a water advisory was issued, was 5.0%,
significantly higher than before the source water was changed to Flint River water (April 2013-April
2014, a proportion of 3.1%). Multivariate adjusted odds ratios comparing the odds of EBLLs were 1.46
(95% CI: 1.06, 2.01), 1.28 (95% CI: 0.92, 1.76), and 0.75 (95% CI: 0.51, 1.12) for the period afterthe
switch to Flint River water and before the water advisory, after the switch to Flint River water and after
the water advisory, and after the switch back to Lake Huron water, respectively.
Over a longer period of time, BLLs of children in Flint, MI were found to decrease, which is
consistent with national trends. Gomez et al. (2019) compared BLLs of children <5 years old from the
periods of April 2006-October 2007, April 2012-October 2013, and April 2014-October 2015, finding
GMs of BLLs decreased from 2.19 ± 0.03 (ig/dL to 1.47 ± 0.02 (ig/dL and finally to 1.32 ± 0.02 (ig/dL,
respectively. In addition, Gomez et al. (2019) found that among a population of women 12-50 years of
age, GMs decreased from the period of April 2012-October 2013 (0.69 (ig/dL; 95% CI 0.63, 0.75) to
April 2014-October 2015 (0.65 ^ig/dL; 95% CI: 0.60, 0.71) to April 2016-October 2017 (0.55 ^ig/dL;
95% CI 0.54, 0.56).
Research has shown that Pb can be transferred between individuals through blood transfusions.
Elabiad and Hook (2013) investigated Pb concentrations in 322 transfusions given to low-birth-weight
infants. The average Pb level found in each packed RBC unit was 18.3 ±1.3 |ig/kg. and the average Pb
load from each transfusion was 0.21 ± 0.13 |ig/kg. Gehrie et al. (2013) found a median Pb concentration
of 0.8 (ig/dL with a SD of 0.80 (ig/dL among 100 packed RBC units.
2.4.2 Pb in Bone
The 2013 Pb ISA (U.S. EPA. 2013) provides a detailed list of studies going back to 1994 that
contain bone measurements of Pb. In nonoccupationally exposed individuals, typical group mean tibia
bone Pb concentrations ranged from 10 to 30 |ig/g. Bone Pb data for occupationally exposed individuals
were also generally higher compared with nonoccupationally exposed individuals. The literature search
and screening revealed only a few studies related to Pb concentrations in bone for this current document.
NHANES does not contain bone Pb concentrations, so this information must be retrieved from studies in
the literature.
Wilkeretal. (2011) investigated Pb concentration changes over time in the mid-tibia shaft and
patella bones for subjects in the VA Normative Aging Study between June 1991 and December 2002.
Subjects attended four visits to have bone concentrations measured using K-XRF with a drop-off in the
number of subjects occurring between visits (n = 554 for 1st tibia measurement, n = 553 for 1st patella
measurement versus n = 73 for 4th tibia measurement, n = 72 for 4th patella measurement). Participants
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had a mean patella Pb measurement of 31.1 (ig/g (SD = 19.9) and a mean tibia Pb measurement of
21.6 (ig/g (SD = 13.6) at the 1st visit. Overall, after adjusting for age at baseline, BMI, years of education,
pack-years smoked, alcoholic drinks/day, instrument used, and vitamin C intake, tibia Pb concentrations
had a decline of 1.4% per year and patella Pb had a decline of 5.1% per year until after 4.6 years when
there was no predicted significant change in patella Pb. Older individuals were found to have higher bone
Pb concentrations.
McNeill et al. (2018) measured bone Pb levels using in vivo XRF in a Toronto, Ontario, Canada
population between 2009 and 2011 and compared them against Hamilton, Ontario, Canada in vivo XRF
measurements of bone Pb collected in the early 1990s. Both groups had no record of occupational
exposure, and home postal code information revealed there was some overlap between recruitment areas
for both studies. The slope of the tibia Pb content versus age was reduced by 36%-56% compared with
17 years prior, showing it is likely that over time, there have been reductions in uptake of Pb into the
bones, from environmental exposures, among the population within the Ontario region.
2.4.3 Pb in Urine
Urine-Pb concentrations for the U.S. population are monitored in NHANES. Data from the most
recent CDC report CDC (2021a) on NHANES data can be found in Table 2-12. NHANES IV data
presented in the 2006 Pb AQCD (U.S. EPA, 2006), 1999-2008 NHANES data presented in the 2013 Pb
ISA (U.S. EPA, 2013), and Table 2-12, show urine |ig Pb/g creatine GM concentrations have continued
to drop overtime, similar to BLLs. As an example, the urine GM Pb concentration for subjects >20 years
of age in the NHANES IV 1999-2000 data was 0.72 (95% CI: 0.68, 0.76), whereas in 2015-2016, it was
0.304 (95% CI: 0.276, 0.315). A discussion of urinary Pb elimination is provided in Section 2.2.3.
Table 2-12
Urine-Pb concentrations in the U.S. population
Survey Stratum
Period
Geometric Mean
(Hg Pb/g CR)a
95% Confidence Interval
Number of
Subjects
All
2011-2012
0.360
0.328, 0.396
2,504
2013-2014
0.277
0.257, 0.298
2,664
2015-2016
0.284
0.261, 0.308
3,061
6-11 yr
2011-2012
0.346
0.292, 0.410
399
2013-2014
0.222
0.192, 0.258
402
2015-2016
0.257
0.238, 0.276
379
12-19 yr
2011-2012
0.259
0.219, 0.305
390
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Survey Stratum
Period
Geometric Mean
(Hg Pb/g CR)a
95% Confidence Interval
Number of
Subjects
2013-2014
0.201
0.166,
0.245
451
2015-2016
0.196
0.183,
0.211
402
>20 yr
2011-2012
0.381
0.348,
0.416
1,715
2013-2014
0.297
0.280,
0.315
1,811
2015-2016
0.304
0.276,
0.334
1,794
Males
2011-2012
0.414
0.367,
0.466
1,262
2013-2014
0.315
0.295,
0.337
1,318
2015-2016
0.313
0.285,
0.343
1,524
Females
2011-2012
0.316
0.282,
0.355
1,242
2013-2014
0.245
0.222,
0.269
1,346
2015-2016
0.259
0.233,
0.288
1,537
Mexican
2011-2012
0.372
0.320,
0.431
317
Americans
2013-2014
0.277
0.240,
0.319
453
2015-2016
0.295
0.260,
0.335
585
Non-Hispanic
2011-2012
0.431
0.385,
0.483
669
Black People
2013-2014
0.371
0.320,
0.429
581
2015-2016
0.340
0.298,
0.388
671
Non-Hispanic
2011-2012
0.346
0.311,
0.385
820
White People
2013-2014
0.267
0.245,
0.290
985
2015-2016
0.275
0.247,
0.305
924
All Hispanic
2011-2012
0.372
0.327,
0.423
573
People
2013-2014
0.270
0.239,
0.305
701
2015-2016
0.284
0.258,
0.312
982
Asian People
2011-2012
0.383
0.341,
0.429
353
2013-2014
0.257
0.230,
0.287
292
2015-2016
0.292
0.264,
0.324
332
yr = year(s).
aValues are in |jg Pb/g creatine (CR). Source: Data sourced from CDC (2021a1.
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2.4.4
Pb in Other Biomarkers
Biomarkers other than blood and bone Pb have been used in various studies to measure Pb body
burden, although they are not as well established. Robbins et al. (2010) analyzed tooth enamel samples
from 127 individuals born between 1936 and 1993 and found the log-transform of tooth enamel
concentration was significantly predicted by the log-transform of Lake Erie sediment core data (i.e., Pb
concentrations found in the Lake Erie sediment) obtained by Granev et al. (1995) (p < 0.00001) and by
the log-transform of U.S. consumption of Pb in gasoline (p < 0.00001). Studies performed in Brazil found
Pb concentrations in tooth enamel among 4- to 6-year-old kindergarteners in Sao Paulo to be significantly
higher (p < 0.0001) for those living near a Pb-acid battery processing plant than those living in other parts
of the city (control versus exposed medians: 206 mg/kg versus 786 mg/kg) and Pb in tooth samples to be
higher for children 4-12 years of age living near a dam with heavy metal sediments compared with
children 4-13 years of age living in a control area (control versus exposed averages: 0.91 mg/kg versus
1.28 mg/kg) (Arruda-Neto et al.. 2009; de Almeida et al.. 2007). In a study of Pb concentrations in the
general population, Arruda-Neto et al. (2010) observed 10-year-olds had the highest Pb teeth
concentrations, and tooth Pb concentrations stayed constant in adulthood but dropped to just above 30%
among 64-year-old subjects, although they did not adjust for confounding factors. Johnston et al. (2019)
measured Pb concentrations in 50 deciduous teeth of 43 children living in Los Angeles and modeled soil
Pb concentrations in the area using data from the California Department of Toxic Substances Control. The
authors found mean prenatal Pb concentrations, reported as 2ll8Pb:43Ca, were 4.104 x 10 4. and the mean
postnatal level was 4.109 * 10 4. Soil Pb exposure was a predictor of teeth Pb concentrations.
Jursa et al. (2018) measured Pb concentration levels in the hair of 222 children in the Mid-Ohio
Valley region. The median Pb concentration was 0.15 (ig/g, with Pb levels higher in males than in females
and varying along hair length. Sears et al. (2012) performed a systematic review of Pb secretion in sweat,
finding eleven studies. Sweat concentrations were found to vary considerably across studies, with sweat
Pb levels up to 283 |ig/L in nonoccupationally-exposed subjects. There was mixed evidence as to whether
secretion of Pb through the skin could lower PbB.
2.5 Empirical Models of Pb Exposure-Blood Pb Relationships
Multivariate regression models, commonly used in epidemiology, provide estimates of the
variability in BLL (or other biomarker) potentially explained by various exposure pathways (e.g., air Pb
concentration, surface dust-Pb concentration). Structural equation modeling links several regression
models together to estimate the influence of determinants on the internal dose metric. Regression models
can provide estimates of a change of blood or bone Pb concentration in response to an incremental change
in exposure level (i.e., slope factor). One strength of regression models for this purpose is that they are
empirically verified within the domain of observation and have quantitative estimates of uncertainty
embedded in the model structure. However, regression models are based on (and require) paired
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predictor-outcome data and, therefore, the resulting predictions are confined to the domain of
observations and are typically not generalizable to other populations. Regression models also frequently
exclude numerous parameters that are known to influence human Pb exposures (e.g., soil and dust
ingestion rates) and the relationship between human exposure and tissue Pb levels, parameters that are
expected to vary spatially and temporally. Thus, extrapolation of regression models to other spatial or
temporal contexts can be problematic.
A variety of factors may potentially affect estimates of blood Pb-air Pb slope factors.
Simultaneous changes in other (non-air) sources of Pb exposure can affect the relationship indicated for
air Pb. For example, remedial programs (e.g., community and home-based dust control and education)
may be responsible for partial blood Pb reduction seen in some studies. The effect of remedial programs
may lead to an overestimation of declines in blood Pb due to changes in air Pb and a corresponding
positive bias in blood Pb-air Pb slopes. However, model adjustment for remedial programs and other
factors (e.g., soil Pb concentrations) may also cause a negative bias in blood Pb-air Pb slopes. A tendency
over time for children with lower BLLs to not return for follow-up testing has been reported. The follow-
up of children with higher BLLs would likely lead to an underestimation of reductions in blood Pb
following reductions in air Pb and cause a negative bias in blood Pb-air Pb slopes. Another factor is the
extent to which all the air Pb exposure pathways are captured by the data set and its analysis. For
example, some pathways (such as exposure through the diet or surface soils) may respond more slowly to
changes in air Pb than others (such as inhalation). Additionally, some studies may include adjustments for
variables that also reflect an influence from air Pb (e.g., SES or soil Pb). With air Pb concentrations
decreasing overtime, remaining Pb sources (including contributions of legacy airborne Pb to soil and
dust) may be the dominant contributors to current BLLs. Not accounting for Pb exposure from sources
other than current air Pb may positively bias estimates of the influence of current air Pb concentrations
(PbA) on blood Pb. Studies may also vary in the ages of subjects, which given age-related changes in
blood Pb can also influence estimates.
Many studies have used TSP measurements of PbA. The sampling efficiency of TSP samplers is
affected by particle size distribution, wind speed, and wind direction. For example, especially for larger
particles (aerodynamic diameter >20 |im). TSP sampling efficiency decreases with increasing wind speed
(see Appendix 1). Such effects on TSP sampling efficiency can, in areas where such large particles are a
substantial portion of airborne Pb, lead to uncertainties in the comparability of PbA between samples
within a study and across studies. A uniformly low bias in PbA in a study could positively bias estimated
blood Pb-air Pb slopes for that study. Moreover, variability in TSP samples is likely to result from
temporal variation in wind speed, wind direction, and source strength. Such temporal variability would
tend to increase uncertainty and reduce the statistical strength of the relationship between air Pb and blood
Pb but may not necessarily affect the slope of this relationship. A number of factors, including those
described above, cause uncertainty in the magnitude of estimated blood Pb-air Pb slope factors and may
lead to both positive and negative biases in the estimates from individual studies.
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2.5.1
Air Pb-Blood Pb Relationships in Children
Within the literature and U.S. EPA documents, the relationship between air Pb and blood Pb is
commonly characterized in terms a "slope factor" or "air-to-blood ratio." An air-to-blood ratio of 1:5
indicates that for every 1 (ig/m3 of air Pb, there is a 5 (ig/dL increase in blood Pb. Synonymously, this is
characterized by a slope factor of 5 (ig/dL per |ig/nr\ The 1986 Pb AQCD (U.S. EPA, 1986) described
epidemiologic studies of relationships between air Pb and blood Pb. Drawing from the studies examined,
the aggregate blood Pb-air Pb slope factor (when considering both air Pb and Pb in other media derived
from air Pb) was estimated to be approximately double the slope estimated from the contribution due to
inhaled air alone (U.S. EPA, 1986). Much of the pertinent earlier literature (e.g., prior to 1984, when air
Pb was dominated by the use of leaded gasoline in on-road motor vehicles) on children's BLLs was
summarized by Brunekreef (1984). The 1986 Pb AQCD also noted ratios derived from occupational
studies of adult cohorts involving higher blood and air Pb levels are generally smaller than ratios from
population studies involving lower blood and air Pb levels [see the 1986 Pb AQCD, Chapter 11, p. 99
(U.S. EPA. 1986)1. Most studies have empirically modeled the air Pb to blood Pb relationship using
nonlinear regression (i.e., log-log), which itself gives an increasing slope with decreasing air Pb
concentration. In the 2008 final rule for the Pb NAAQS (73 FR 66964), the U.S. EPA, recognizing
uncertainty and variability in the air-to-blood relationships, interpreted the evidence as providing support
for a range of estimates inclusive of 1:5 at the lower end and 1:10 at the upper end, with the ratio of 1:7
identified as a central estimate within the range supported by the evidence at the time (73 FR
67001-67002, 67005).
At the time of the 2013 Pb ISA (U.S. EPA, 2013), due to the limited evidence, there was
uncertainty in projecting the magnitude of the air Pb-blood Pb relationship to ambient PbA below
0.2 |ig/m\ The air Pb-blood Pb relationship in terms of slope factors or ratios was not discussed in the
2006 Pb AQCD (U.S. EPA, 2006). There are studies since the 2013 Pb ISA that evaluate the air Pb-to-
blood Pb relationship that are more reflective of current conditions with central tendency PbA between
0.004 and 0.04 |ig/nr\ Table 2-13 summarizes new and old studies from which air-to-blood ratios were
derived. With the exception of Ranft et al. (2008), slopes corresponding to a central estimate of the PbA
of each study are provided. Slope factor data as a function of PbA are illustrated in Figure 2-16, which
shows slope factors continue to increase with decreasing PbA seen in the newer studies. Although
saturable GI absorption and saturation of Pb binding to RBC occur at relatively high rates of Pb intake
leading to PbB of 20-30 (ig/dL (see Section 2.2), the nonlinear relationship between PbA and PbB cannot
be explained by abiokinetic mechanism. With reference to Table 2-13 and Figure 2-16, it is readily
apparent that the PbA are considerably higher in older studies (Hilts, 2003; Tripathi et al„ 2001; Hayes et
al„ 1994; Schwartz and Pitcher, 1989; Brunekreef, 1984) than newer studies (Mcng et al., 2014;
Richmond-Bryant et al., 2014; Richmond-Bryant et al., 2013; Zahran et al., 2013a; Bierkens et al„ 2011).
In general, longitudinal studies conducted after phasing out leaded gasoline would best inform the
current relationship of PbB change corresponding to PbA fluctuation. Ideally, such studies would
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compare two populations for which all Pb sources are relatively constant with only changes in PbA
concentrations. Such a nearly ideal study, Hilts (2003) reported the change in PbB from 1996 to 2001 for
children under five years old associated with the emission reduction from a local smelter in Trail, BC,
Canada. However, even in this study, the reduction in exposure from pathways other than air cannot be
ruled out due to the "comprehensive education and case management programs/' Common to all studies,
Pb in other media, not just air, were also decreasing, e.g., due to stopping use of Pb solder in food cans
and plumbing. An advancement in analyses of PbB-PbA associations came from leveraging the U.S. EPA
AQS with NHANES surveys. The PbB-PbA associations across different NHANES periods should reflect
the change in this association for the U.S. population overtime (Richmond-Bryant et al.. 2014;
Richmond-Bryant et al.. 2013) because each NHANES is a representative sample of the U.S. population.
However, merging PbB results from multiple NHANES periods with the U.S. EPA AQS could introduce
exposure measurement errors as well as uncertainties in terms of population representativeness and
availability of covariates. Each single study presented in Table 2-13 deviates from the ideal design in one
or more aspects. Collectively, all of these studies contribute to our understanding of how PbA impacts
PbB.
For log-log relationships between total blood Pb and PbA, the instantaneous slope factor,
d[PbB]/d[PbA], was by Equation 2-2.
= bPo x fjpbA X [PbA](Pp"A~r) Equation 2-2
where: b is the logarithm base (either 2.7183 or 10) used in a study, bO is the regression intercept,
bPbA is the regression slope, and PbA is central estimate of air Pb for the study. The instantaneous slope
calculated by Equation 2-2 provides the same estimated slope as derived by evaluating regression
equations at ±0.01 |ig/m3 from central estimate of air Pb as done for studies in Table 3-12 of the 2013 Pb
ISA (U.S. EPA. 2013). New and old studies vary with regard to the use of single or multivariate
regression and, for the latter, with regard to the variables included. Newer studies (Meng et al.. 2014;
Richmond-Bryant et al.. 2014; Richmond-Bryant et al.. 2013; Zahran et al.. 2013a; Bierkens et al.. 2011)
that provide estimates for a total blood Pb-air Pb slope factor are described below. The series of studies
by Meng et al. and Richmond-Bryant et al. were conducted by U.S. EPA to address slope factor-related
uncertainties identified while completing the 2013 Pb ISA (U.S. EPA. 2013). Older studies were
discussed in detail in Section 3.5.1 of the 2013 Pb ISA (U.S. EPA. 2013).
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Table 2-13 Summary of estimated slopes for blood Pb-to-air Pb slope factors
in children
Reference
Study Methods
Model Description
Blood Pb-Air Pb Slope3
Child Populations-Air
Location: European countries
Yr: 1999-2008
Bierkens et al Subjects: Children (<6 yr; n = 28)
(2011) Analysis: Univariate regression of
blood Pb from literature and air Pb
from a European Environment
Agency database
Model: Log-Log
Blood Pb: 1.45-4.11 |jg/dL (mean
range for study groups)
Air Pb: 0.001-0.056 |jg/m3 (annual
mean range for study groups)
12.0 (0.020)b
Brunekreef
(1984)
Location: Various countries
Yr: 1974-1983
Subjects: Children (varying age
groups including children from 0 to
18 yr; n > 190,000)
Analysis: Meta-analysis of 96 child
populations from 18 study locations
Model: Log-Log
Blood Pb: 5-76 |jg/dL (mean range
for study populations)
Air Pb: 0.1-10.0 |jg/m3 (mean range
for study locations, averaging time
not typically indicated)
All children: 4.6 (1 5)c
Children <20 ugldL: 4.8 (0.54)d
Haves et al.
(1994)
Location: Chicago, IL
Yr: 1974-1988
Subjects: 0.5-5 yr (n = 9,604)
Analysis: Regression of quarterly
median blood Pb and quarterly mean
air Pb
Model: Log-Log
Blood Pb: 10-28 |jg/dL (quarterly
median range)
Air Pb: 0.05-1.2 |jg/m3
(quarterly mean range)
8.2 (0.62)e
Location: Trail, BC
Yr: 1996-2001
Subjects: 0.5-5 yr, 1996-2000; 0.5-
3 yr, 2001 (Estimated n = 220-460
Hilts (2003) Per yr' based on 292-536 eligible
children per yr with 75%-85%
participation)
Analysis: Regression of blood Pb
screening and community air Pb
following upgrading of a local smelter
Model: Linear
Blood Pb: 4.71-1.5 |jg/dL (annual
GM range)
Air Pb: 0.13-1.1 |jg/m3 (annual GM
range except 2001, which reflects a
9-mo average)
7.0 (0.48)f
Location: United States (contiguous
states)
Yr: 1999-2008
Subjects: 1-5 yr (n = 178, TSP;
Menq et al. n = 2,150, PM10), 6-11 yr (n = 212,
(2014) TSP; n = 2,261, PM-io)
Analysis: Age-stratified linear mixed
effects models were run to assess
the relationship of PbB with PbA
without covariates
Model: Log-Log
Blood Pb: 1.7-2.3 |jg/dL and 1.4-
2.0 |jg/dL(range of GMs),
respectively, for 1 - to 11 -yr-old
children paired with TSP and PM10
data
Air Pb: 0.0135-0.0151 |jg/m3and
0.0051-0.0054 |jg/m3 (range of GMs
of daily PbA in TSP and PM10,
respectively, paired with BLL data for
1- to 11-yr-old children)
1-5 yr
9.1 (TSP, 0.0135)8
37.7 (PM10, 0.0054)
6-11 yr
3.0 (TSP, 0.0151)
20.1 (PM10, 0.0051)
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Reference
Study Methods
Model Description
Blood Pb-Air Pb Slope3
Schwartz and
Pitcher (1989):
U.S. EPA (1986)
Location: Chicago, IL
Yr: 1976-1980
Subjects: Black children, 0-5 yr
(n = 5,476)
Analysis: Multivariate regression of
blood Pb with mass of Pb in gasoline
(derived from gasoline consumption
data and Pb concentrations in
gasoline for the United States)
Model: Linear
Blood Pb: 18-27 |jg/dL(mean
range)h
Air Pb: 0.36-1.22 |jg/m3 (annual
maximum quarterly mean)'
3.6 (0.75)'
Location: Mumbai, India (multiple
residential locations)
T. x , Yr: 1984-1996
Tripathi et al.
(2ooi ^ Subjects: 6-10 yr (n = 544)
Analysis: Regression of residential
location-specific average blood Pb
and air Pb data
Model: Linear
Blood Pb: 8.61-4.4 |jg/dL (GM
range for residential locations)
Air Pb: 0.10-1.18 |jg/m3 (GM range
of 24-hour samples at residential
locations)
Richmond-
Brvant et al.
(2014):
Richmond-
Brvant et al.
(2013)
Location: United States (contiguous
states)
Yr: 1988-1994, 1999-2008
Subjects: 1-5 yr (n = 759), 6-11 yr
(n = 516)
Analysis: Age-stratified linear mixed
effects models were run to assess
the relationship of PbB with PbA,
with and without covariates (age,
household size, mother's age,
poverty-income ratio, and street
length)
Model: Log-Log
Blood Pb: 1.7-4.5 |jg/dL (range of
medians among surveys and ages)
Air Pb: 0.011-0.037 |jg/m3 (range of
median annual averages among
surveys and ages)
1-5 yr
16.4 (0.037)'
15.3 (0.011)
6-11 yr
15.7 (0.036)
16.5 (0.016)
Zahran et al.
(2013a)
Location: Detroit, Ml
Yr: 2001-2009
Subjects: 0-<1 yr (n = 19,265), 1-
<2 yr (n = 75,070), 2-<3 yr
(n = 58,500), 3-<4 yr (n = 66,507),
4-<5 yr (n = 67,061), 5-<6 yr
(n = 34,073), 6-<7 yr (n = 18,911),
7-<8 yr (n = 8,649), 8-<11 yr
(n = 13,610)
Analysis: Age-stratified fixed effect
regression controlling for
confounding variables (Pb facility,
capillary blood draw, sex, yr,
meteorology)
Model: Log-Log
Blood Pb: <5 |jg/dL (67%), >5 |jg/dL
(33%)
Air Pb: 0.004 ± 0.001 |jg/m3
(monthly mean ± SD)
0 yr: 34.0 (0.004)m
1 yr: 57.3 (0.004)
2 yr: 62.5 (0.004)
3 yr: 37.2 (0.004)
4 yr: 30.9 (0.004)
5 yr: 35.5 (0.004)
6 yr: 24.8 (0.004)
7 yr: 21.3 (0.004)
8-10 yr: 16.7 (0.004)
Child Populations - Air and Soil
Ranft et al.
(2008)
Location: Germany
Yr: 1983-2000 (blood Pb and air
Pb), 2000-2001 (soil Pb)n
Subjects: 6-11 yr (n = 843)
Analysis: Pooled multivariate
regression of five cross-sectional
studies
Model: Log-Linear
Blood Pb: 2.21-3.6 |jg/dL (5th-
95th percentile)
Air Pb: 0.03-0.47 |jg/m3 (5th-
95th percentile of annual average)
3.2, 6.4°
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Reference
Study Methods
Model Description
Blood Pb-Air Pb Slope3
Mixed Child-Adult Populations
Schwartz and
Pitcher (1989);
U.S. EPA (1986)
Location: United States
Yr: 1976-1980
Subjects: NHANES II, 0.5-74 yr,
whites (n = 9,987)
Analysis: Multivariate regression of
blood Pb with mass of Pb in gasoline
(derived from gasoline consumption
data and Pb concentrations in
gasoline for the United States)
Model: Linear
Blood Pb: 11-18 |jg/dL'(mean
range)h
Air Pb: 0.36-1.22 |jg/m3 (annual
maximum quarterly mean)'
9.3 (0.75)p
BLL = blood lead level; GM = geometric mean; GSD = geometric standard deviation; mo = month; NHANES = National Health and
Nutrition Examination Survey; PbA = air Pb concentration (|jg/m3); PbB = blood Pb concentration (|jg/dL); PM = particulate matter;
TSP = total suspended particles; yr = year(s).
aSlope is predicted change in blood Pb (|jg/dL per |jg/m3) at central estimate of air Pb for the study (shown in parentheses), except
for Ranft et al. (2008) in which the slope from the paper was used because a regression equation was not available. The central
estimate for Brunekreef (1984) and Richmond-Bryant et al. (2014; 2013) was the median of air Pb concentrations, the central
estimate for Meng et al. (2014) was the GM of air Pb concentrations, and for all other studies the mean was used. For multiple
regression models, the slope factor was based only on air Pb coefficient and intercept. Depending on the extent to which other
variables modeled also represent air Pb, this method may underestimate the slope attributable to air pathways. In single
regression models, the extent to which nonmodeled factors, unrelated to air Pb exposures, exert an impact on blood Pb that
covaries with air Pb may lead to the slope presented here to overrepresent the role of air Pb.
blog(PbB) = log(PbA) * 0.09 + 0.58.
cln(PbB) = In(PbA) * 0.3485 + 2.853.
dln(PbB) = In(PbA) * 0.2159 + 2.620.
eln(PbB) = In(PbA) * 0.24 + 3.17.
fPbB = PbA x 7.0.
91-5 years [TSP, In(PbB) = ln(Pb A) * 0.056 + 1.024; PM10, In(PbB) = ln(Pb A) * 0.104+ 1.213]; 6-11 years [TSP, In(PbB) = ln(Pb
A) x 0.028 + 0.596; PM10, In(PbB) = ln(Pb A) * 0.073 + 0.725],
hObserved blood Pb values not provided; data are for regressed adjusted blood Pb.
ipbB = PbA x 8.6.
'Based on air Pb data for United States (1986 Pb AQCD) as a surrogate for Chicago.
kPbB = PbA x 3.6.
'1-5 years [1988-1994, In(PbB) = ln(Pb A) x 0.1395+ 1.9315; 1999-2008, In(PbB) = In(PbA) x 0.0755 + 1.1419]; 6-11 years
[1988-1994, In(PbB) = ln(Pb A) x 0.1535+ 1.8118; 1999-2008, In(PbB) = In(PbA) x 0.1552 + 1.1709],
m0 years [In(PbB) = ln(Pb A) x 0.080 + 0.973]; 1 years [In(PbB) = ln(Pb A) x 0.087 + 1.449]; 2 years [In(PbB) = ln(Pb
A) x 0.069 + 1.669]; 3 years [In(PbB) = ln(Pb A) x 0.040 + 1.535]; 4 years [In(PbB) = ln(Pb A) x 0.036 + 1.432]; 5 years
[In(PbB) = ln(Pb A) x 0.043 + 1.431 ]; 6 years [In(PbB) = ln(Pb A) x 0.031 + 1.333]; 7 years [In(PbB) = ln(Pb A) x 0.026 + 1.331 ]; 8-
10 years [In(PbB) = ln(Pb A) x 0.023 + 1.192]; the mean Pb air concentration was provided by authors to U.S. EPA as a correction
to their paper.
"Study that considered air Pb and soil Pb, wherein the air Pb-blood Pb relationship was adjusted for soil Pb.
"Slope provided in paper with background blood Pb level of 1.5 and 3 |jg/dL, respectively, and a GM blood Pb ratio of 2.55 for
ambient air.
"PbB = PbA x 9.63.
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100 T
tu>
2.
-------
models, the estimated effect (i.e., the regression slope, PpbA) was higher for the earlier time period. In
addition, when the effect estimates calculated in this study were compared with values reported in older
studies (Bierkens et al.. 2011; Haves et al.. 1994; Brunekreef. 1984). a declining trend was observed,
indicating a decrease in the influence of PbA on PbB over this time. The authors also note that for young
children (1-5 years), the effect estimate for 1999-2008 data decreased when models included covariate
factors, whereas the effect estimate for the 1988-1994 data was not significantly different between
adjusted and unadjusted models. The authors suggested this finding may indicate that estimates of PbA on
blood Pb may have a positive bias when not corrected for covariates, and this inflation may be more
apparent at lower air PbA concentrations. In Richmond-Bryant et al. (2014). slope factors were estimated
by Equation 2-2 and compared with other published data. In Table 2-13 and Figure 2-16, the slope factors
were calculated using the median blood Pb of children because they were reported for age grouping (1-5
and 6-11 years) and NHANES survey period (1998-1994 and 1999-2008), whereas mean blood Pb was
not reported. The authors concluded their NHANES regression results, compared with those from the
literature, show the slope factor increases with decreasing air Pb among children 0-11 years of age.
Meng et al. (2014) merged participant-level data for blood Pb from 1999-2008 NHANES with air
Pb data for TSP, PMi0, and PM2.5 from the U.S. EPA AQS. A 4-km neighborhood scale was used to
represent PbA concentrations in urban areas (not near sources) to merge with NHANES data. This was
the first (and currently only) study comparing the relationship between blood Pb and airborne Pb among
multiple size fractionated PM samplers rather than TSP samplers only. The impetus for this research was,
in part, due to another U.S. EPA study (Cho et al.. 2011) showing the mass median diameter of airborne
Pb had shifted from <2.5 (mi prior to the phase-out of leaded gasoline to somewhere between 2.5 and
10 (mi after the phase-out, which might alter exposure pathways and PbA-PbB relationships. They
examined the relationship between PbA and PbB by particle size (TSP, PM10, and PM2.5) and by age
groups included in NHANES sample design (only children <12 years are presented in Table 2-13). PbA
in PM10 was significantly (p < 0.01) related to PbB for all age groups. While PbA in TSP was
significantly (p <0.05) related to PbB for the 12-19 and 20-59 age groups, it was not statistically
significant for children <12 years of age for Pb-TSP. However, it is provided in Table 2-13 and
Figure 2-16 for comparison with other recent studies (Richmond-Bryant et al.. 2014; Richmond-Bryant et
al.. 2013) that used Pb-TSP. This study also found a positive association of Pb-PlVbs (p <0.05) with BLLs
of children in the 6-11 age group but there was a lack of a significant relationship for other age groups.
The lack of a significant relationship for PM2.5 may, in part, be attributed to airborne Pb being found
associated with particles larger than 2.5 (mi (Cho et al.. 2011) during the 1999-2008 period. However, the
authors also note the data for PM2.5 are inherently more uncertain than the other air Pb measurements used
because a large portion (~60%) of the PM2.5 PbA were below detection limits compared with TSP (25%)
and PM10 (34%). In addition, it should be noted that sample sizes are very different among PM2 5 (n = 193
for children aged 1-5 years) PM10 (n = 2150 for children aged 1-5 years) and TSP (n = 178 for children
aged 1-5 years), and that this is a potential factor affecting statistical significance. To derive slope factors
at the central tendency for air Pb, it was necessary to solve for p0 (see Equation 2-2) using data from
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Tables 2 and 3 of Meng et al. (2014). It was also necessary to use the GM air Pb from Table 1 of Mcng ct
al. (2014) to calculate slope factors as the mean was not available. The GM also showed more variation
between age groups than the median and thus was assumed more representative of the central tendency.
An important finding from this study was that blood Pb was more consistently and strongly associated
with PMio than either TSP or PM2.5, which may relate to sample size as noted above.
Richmond-Bryant et al. (2015) provided the first study assessing effect modification of age, sex,
housing age, and race/ethnicity on the relationship between blood Pb and air Pb. The authors used merged
participant-level data for blood Pb from 1999-2008 NHANES with U.S. EPA AQS Pb-PMio data because
their prior work (Meng et al.. 2014) showed Pb-PMio data had the strongest associations with blood Pb.
Consistent with their prior studies, the authors merged NHANES data with 4-km neighborhood scale PbA
concentrations to represent urban areas not near sources. Effect estimates (i.e., the regression slope, PpbA)
were higher for children (1-5, 6-11, and 12-19 years) than for adults or all ages. Living in pre-1950
housing contributed to a higher effect estimate for 1- to 5-year-old children, but not for older ages.
Zahran et al. (2013a) examined the association between children's blood Pb and Pb in air and
suspended soil in Detroit, MI using data acquired from January 2001 to December 2009. Estimates for
resuspended soil concentrations were derived from measurements of airborne elements known to
originate from soil at specific ratios (aluminum, silica, calcium, iron, and titanium). Measurements for
airborne Pb (TSP) and elements used to calculate atmospheric soil concentrations (IMPROVE monitor
sampling; PM2 5) were obtained for the Detroit metropolitan area. Blood Pb data were obtained from the
Michigan Department of Community Health. Concentrations of both Pb and resuspended soil in air were
highest in June-September of each year, peaking in August. Air Pb and resuspended soil were 1.45-times
and 1.62-times higher, respectively, in August relative to January. Children's blood Pb was also elevated
in July-September with peaks in July and August that were 1.13-times (95% CI: 1.12, 1.14) greater than
in January. The authors' analyses showed daily variation in air Pb was associated statistically with daily
variation in resuspended soil, suggesting resuspended soil is the major source of urban air Pb. However,
they note the effect may be more significant in younger children. Their model found a standard deviation
rise in atmospheric Pb is associated with a 0.232 (ig/dL (95% CI: 0.203 to 0.26 (ig/dL) increase in
monthly average BLLs of children 0-2 years old compared with a 0.152 (ig/dL (95% CI: 0.13 to
0.173 (ig/dL) increase in children >6 years old. They note this outcome is consistent with prior research
and attribute this to higher exposure in younger children through ingestion of fine particles during hand-
to-mouth contact. Richmond-Bryant et al. (2014) reported a slope factor of -60 (ig/dL per (ig/m3 for 1- to
2-year-old children in this study, which is the largest illustrated in Figure 2-16.
While the results of the Zahran et al. (2013a) study demonstrated children's BLLs in Detroit
varied with season, attributing this variability solely to air Pb does not account for changes to children's
BLLs resulting from seasonal changes in Pb in other exposure media. For example, water Pb
concentrations in flushed water samples have been observed to be increased in the summer relative to the
winter (Ngueta et al.. 2014; Deshommes et al.. 2013). Although magnitude of water Pb concentrations
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was considerably greater in homes with Pb-service lines than without, the water Pb concentrations were
five to six times greater in July relative to December in both cases, based on Figure 1 of Ngueta et al.
(2014). Recent modeling of soil/dust ingestion rates by Ozkavnak et al. (2022) found mean daily soil
ingestion rates to be approximately doubled (based on Tables S7-S10 of the paper) in the summer relative
to the rest of the seasons by increasing from 8 to 15 mg/day in 1- to <2-year-olds, 20 to 47 mg/day in 2-
to <3-year-olds, and from 23 to 57 mg/day in 3- to <6-year-olds. Consider the effects on blood Pb from
modest soil Pb (no dust) and water Pb concentrations of 50 ppm (i.e., mg/kg) and 0.9 ppb (i.e., |ig/L).
respectively. To simulate season effects using IEUBK v2.0 for 1- to <6-year-old children, the default soil
ingestion rate can be lowered by 25% for winter (relative to an annual rate), increasing the default soil
ingestion rate by 170% to simulate summer (relative to an annual rate), and increasing the water Pb
concentration by five times to simulate effects of season (i.e., 0.9 ppm in winter and 4.5 ppm in summer).
This IEUBK v2.0 simulation predicts a winter-to-summer increase in the GM BLL of 0.7 (ig/dL in 1- to
<3-year-olds and 0.6 (ig/dL in 3- to <6-year-olds. Figure SI1 ofZahran et al. (2013a) shows median
seasonal change in blood Pb of -0.8 (ig/dL across all years of the study. Unaccounted for seasonal effects
on water Pb and soil ingestion rates could account for most of the blood Pb changes attributed to air Pb by
Zahran et al. (2013a).
2.5.2 Air Pb-Blood Pb Relationships in Adults
2.5.2.1 General Populations
Several of the new publications since the last Pb ISA (U.S. EPA. 2013) provide estimates of slope
factors for both children and adults. As the methods of these new publications are discussed above in
Section 2.5.1, only the results for adults contrasted with those for children are provided here. Bierkens et
al. (2011) analyzed the blood Pb - air Pb relationship in 174 adults (+18 years; 1981-2008; 70 males, 84
females, 20 unspecified). Regression results for women [Log(blood Pb) = 0.79 + 0.34 x Log(Air Pb);
r = 0.3922; p < 0.001] and men [Log B-Pb = 0.97 + 0.44 x Log(Air Pb); r2 = 0.5158; p < 0.001] yielded
respective slope factors of 11 and 17 (ig/dL per (ig/m3 at the central air Pb of 0.076 |ig/nr\ Although the
central air Pb was higher for these adults (0.076 (ig/m3) than preschool children (0.020 (ig/m3) in this
study, the slope factors were still greater in the adult population. Although Richmond-Bryant et al. (2013)
and Richmond-Bryant et al. (2014) assessed the relationship between air Pb and blood Pb in adults (ages
20-59 years and >60 years) for the 1988-1994 and 1999-2008 periods, intercepts were not reported for
the calculation of slope factors. However, the effect estimate PpbA was similar in magnitude between
children and 20- to 59-year-old adults. For >60-year-old adults, the effect estimate PpbA was also similar to
children in the 1988-1994 period, whereas PpbA was negative in the later 1999-2008 period. For the
1999-2008 period and air Pb in PMio, Meng et al. (2014) reported BLLs were more sensitive to the
changes in PbA in children (1-5 and 6-11 years) and older adults (>60 years) than teenagers (12-
19 years) and adults (20-59 years). Slope factors (fig/dL per (ig/m3) estimated for this study by rank are
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38 (1- to 5-year-olds), 30 (>60 years), 20 (6- to 11-year-olds), 11 (12- to 19-year-olds), and 8 (20- to 59-
year-olds). The negative PpbA for older adults (>60 years) during the 1999-2008 period was also observed
by Meng et al. (2014) for TSP but not for PMio data.
2.5.2.2 Occupational Cohorts
At the time of the 1986 Pb AQCD, there was a great deal of information on blood Pb responses to
air Pb exposures of workers in Pb-related occupations (U.S. EPA. 1986). Almost all such exposures were
at air Pb exposures far in excess of typical nonoccupational exposures and usually did not account for
other potential sources of Pb exposure. The air Pb-blood Pb slopes in these studies were generally much
less (i.e., 0.03-0.2; 1986 AQCD, p. 11-106) than those observed in children when considering aggregate
air Pb contributions (i.e., 3-5; 1986 AQCD, p. 11-106). In addition, the PbA in occupational studies are
typically collected at much shorter durations (e.g., over an 8-hour workday) compared with ambient air
Pb monitoring (which generally involves 24-hour samples), making it difficult to draw comparisons
between occupationally and nonoccupationally-exposed populations. Nonoccupational studies remain the
focus of this appendix. Therefore, only a few occupational studies are presented below to demonstrate
that more recent air Pb and BLLs remain much higher in these studies compared with those conducted in
the general population.
Rodrigues et al. (2010) examined factors contributing to variability in blood Pb concentration in
New England bridge painters, who regularly use electric grinders to prepare surfaces for painting. The
study included 84 adults (83 males, 1 female) who were observed during a 2-week period in 1994 or
1995. The GM air Pb concentration obtained from personal PM samplers worn over the workday was
58 (ig/m3 (GSD 2.8), with a maximum daily value of 210 |ig/nr\ Personal air Pb concentrations were
corrected for respirator use. Mean task personal Pb concentrations were divided by the NIOSH-assigned
protection factor for each respirator reported by workers in their daily diary. Hand-wipe samples were
collected and analyzed for Pb (GM = 793 (ig, GSD 3.7). Blood Pb samples were collected at the
beginning of the 2-week period (GM = 16.1 (ig/dL, GSD 1.7; a level substantially above the general
population) and at the end of the period (GM =18.2 (ig/dL, GD = 1.6). Associations between exposure
variables and blood Pb concentrations were explored with multivariate regression models. When the
model excluded hand-wipe data, the regression coefficient for the relationship between ln[blood Pb
concentration (|ig/dL)] and ln[air Pb (|ig/m3)] was 0.11 (SE = 0.05, p = 0.03). This corresponds to a slope
of 0.009 (ig/dL per (ig/m3 at the GM air Pb concentration for the study. A second regression model
included hand-wipe Pb (n = 54) and yielded a regression coefficient of 0.05 (SE = 0.07, p = 0.45), which
corresponds to a slope of 0.02 (ig/dL per (.ig/rn3 at the GM air Pb concentration for the study.
Two other studies examined the air Pb-blood Pb relationship in occupational settings at higher air
Pb concentrations (GM of 82 and 111 (.ig/rn3 for Pb battery and crystal workers, respectively) (Pierre et
al.. 2002; Lai et al.. 1997). BLLs for the Pb battery workers averaged 56.9 (ig/dL (SD 25.3); for the
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crystal workers, it averaged 21.9 (ig/dL. Both studies employed log-log regression models, resulting in
slopes of 0.04 (Pierre et al.. 2002) and 0.09 (Lai et al.. 1997). Workers in Pierre et al. (2002) had lengths
of service that ranged from a mean of 4.8 years (SD 3.5) in the "others" group to a mean of 22 years (SD
10.3) in the "sandstone grinders" group and there is no mention of respiratory protection. Workers in Lai
et al. (1997) had lengths of service that ranged from less than six months (69 workers) to more than seven
years and there was a mix of subjects wearing cotton masks, 43.2% of males and 60.4% of females.
2.5.3 Soil Pb-Blood Pb Relationships
Slope factor models represent empirically based relationships between BLLs and intake of Pb
and/or Pb concentrations in environmental media. Section 3.5.3 of the 2013 Pb ISA (U.S. EPA, 2013)
provides a description of these models and past reviews of environmental Pb-blood Pb data to develop
these relationships. The U.S. EPA Adult Lead Methodology (ALM) is a slope factor model that has had
extensive use in the U.S. EPA Superfund Program for non-residential adult exposures to Pb in soil (U.S.
EPA, 2003b). The ALM uses 50 mg/day as a plausible central tendency daily ingestion rate for non-
residential exposures including soil in indoor dust resulting from non-contact intensive activities. An
appropriate soil ingestion rate for a construction scenario or other soil contact-intensive scenarios is
100 mg/day. Common exposure scenarios include utility and construction workers, youth trespassers
(>7 years of age), and landscaping. For a given soil Pb concentration, the ALM predicts a GM BLL
(NHANES baseline BLL plus contribution due to soil Pb) and uses a recent NHANES population GSD to
estimate the probability that fetal BLLs will exceed 5 (ig/dL. Alternatively, the model can estimate the
soil Pb concentration that allows no more than a 5% chance of fetal BLLs exceeding 5 (ig/dL.
The California Department of Toxic Substances Control has developed the LeadSpread slope
factor model for assessing Pb in soils to which either children or adults are exposed (CA DTSC, 2022).
LeadSpread has been used in California since 1991. The current version 9 can be used to assess
residential exposures of either children or adults, as well as non-residential exposures of adults. Similar to
the IEUBK model discussed in Section 2.6, LeadSpread uses a GSD of 1.6 to estimate the distribution of
predicted BLLs. Central tendency ingestions rates of 80 mg/day for children and 30 mg/day for adults are
used as recommended by U.S. EPA (2017). For children, LeadSpread predicts the soil Pb concentration
that will prevent an incremental increase of 1 (ig/dL at the 90th percentile of the BLL distribution. For
adults, the model estimates the soil Pb concentration protecting against a fetal BLL increase of 1 (ig/dL at
the 90th percentile.
The 2006 Pb AQCD (U.S. EPA. 2006) and 2013 Pb ISA (U.S. EPA. 2013) also explored the
relationship between blood Pb in children and environmental Pb concentrations. Several analyses of
epidemiologic data found soil and dust exposures were significant predictors of blood Pb concentrations.
Mielke and co-authors have published a series of papers (Mielke et al.. 2019b; Zahran et al.. 2011; Mielke
et al.. 2007) demonstrating the importance of soil Pb as a source of children's Pb exposures in New
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Orleans and other cities. The New Orleans data they developed were especially extensive (>5,000 surface
soil samples; >50,000 blood Pb samples) and included multiple time points demonstrating a now
declining pattern of soil Pb concentrations and BLLs. The statistical analyses in these papers fitted a
nonlinear model between soil Pb level (SLL) and BLL, which becomes increasingly steep for SLLs below
100-200 ppm, as shown in Figure 2-17 that was created using data from the Mielke et al. (2019b) study.
The subsequent Figure 2-18 shows the rapid decline in the slope of the soil Pb to blood Pb relationship
that is most apparent at soil Pb concentrations <20 mg/kg.
12.0
10.0 -
J '
_o
Q.
"o 6.0
_o
CQ
C
.2
T3 4.0
-------
1.0 i
CD
Q_
_0 . .
uo W)
_Q
CL 00
~o E
s
O d)
0.8
0.6
0.4
I ¦
Mielke et al. 2007
Mielke et al. 2019 (new data, unflooded)
•- Mielke et al. 2019 (old data}
Zahran et al. 2011
_Q
CL
_ — 0.2
o
0.0
20 40 60
Soil Lead (mg/kg)
80
=-i
100
1.00 T
-------
samples. As a result of this BLL screening approach, blood Pb measurements may not be statistically
independent because multiple BLLs, including combinations of capillary and venous samples or multiple
venous measurements, may be recorded for one child. An additional complexity is that the concentration
used to characterize a BLL as elevated (and requiring follow-up with confirmatory venous samples) was
lowered from 10 to 5 (ig/dL in 2012 and to 3.5 (ig/dL in 2021 (Ruckart et al.. 2021). As a result of this
change, confirmatory venous samples between 5 and 10 (ig/dL are likely to be relatively rare before the
BLRV was established in 2012. Additionally, the LODs for the analytical methods vary overtime,
depending on the sample type and laboratory, but may not be recorded in BLL databases that are used for
public health screening purposes. In practice, an LOD might be recorded as the measured BLL, making it
difficult to distinguish BLLs that are below the LOD from BLLs that are measured at or near the LOD.
Notably, BLLs may be measured using analyzers that have LODs as high as 3.3 (ig/dL, which is in the
range of current health concerns. The median BLLs from Mielke et al. (2019b) show many are reported at
levels that are less than or equal to the anticipated LOD of the methods: 3.0 (ig/dL and 1.0 (ig/dL for the
first and second survey, respectively.
Mielke et al. (2007) and Mielke et al. (2019b) did not account for potential individual-level
confounding. However, bivariate relationships between BLL, SLL, age of housing, and distance from the
post office, which is inversely associated with Pb exposure, were examined in Mielke et al. (2016) and
Egendorf et al. (2021a). Egendorf et al. (2021a) analyzed these SLL and BLL data for New Orleans by
several variables separately, including distance from the city center, residential racial population, and
household income over two time periods, but did not include a multivariate analysis. Table 2 of Egendorf
et al. (2021a) provides median values for Pb and demographic variables by city sector and shows sectors
closest to the urban core had the highest BLLs, highest SLLs, were primarily Black in racial makeup, had
higher population density, and lower income. All of these variables may reflect important factors
influencing BLLs and that there would be important confounding relationships that would limit the
interpretation of univariate modeling of BLLs versus SLLs. In addition, the publications do not provide
information on how the authors selected the most appropriate statistical model to capture the relationship
between SLL and BLL data. However, it should be noted there is some evidence for nonlinearities at low
PbA, as discussed in Section 2.5.1.
2.6 Biokinetic Models of Pb Exposure-Blood Pb Relationships
An alternative to regression models is mechanistic models, which attempt to specify all
parameters needed to describe the mechanisms (or processes) of transfer of Pb from the environment to
human tissues. Such mechanistic models are more complex than regression models; this added
complexity introduces challenges in terms of their mathematical solution and empirical verification.
However, by incorporating parameters that can be expected to vary spatially or temporally, or across
individuals or populations, mechanistic models can be extrapolated to a wide range of exposure scenarios,
including those that may be outside of the domain of paired predictor-outcome data used to develop the
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model. Exposure-intake models, a type of mechanistic model, are highly simplified mathematical
representations of relationships between levels of Pb in environmental media and human Pb intakes
(e.g., |ig Pb ingested per day). These models include parameters representing processes of Pb transfer
between environmental media (e.g., air to surface dust) and to humans, including rates of human contact
with the media and intakes of the media (e.g., g soil ingested per day). Intake-biokinetic models provide
the analogous mathematical representation of relationships between Pb intakes and Pb levels in body
tissues (e.g., blood Pb concentration). Biokinetic models include parameters that represent processes of
Pb transfer (a) from portals of entry into the body and (b) from blood to tissues and excreta. Linked
together, exposure-intake and intake-biokinetic models (i.e., integrated exposure-intake-biokinetic
models) provide an approach for predicting blood Pb concentrations (or Pb concentrations in other
tissues) that corresponds to a specified exposure (medium, concentration, and duration). Detailed
information on exposure and internal dose can be obtained from controlled experiments but almost never
from epidemiologic observations or from public health monitoring programs. Exposure intake-biokinetic
models can provide these predictions in the absence of complete information on the exposure history and
blood Pb concentrations for an individual (or population) of interest. Therefore, these models are critical
for applying epidemiologic-based information on blood Pb-response relationships to the quantification
and characterization of human health risk. These models are also critical for assessing the potential
impacts of public health programs directed at mitigating Pb exposure or remediating contaminated sites.
However, these models are not without their limitations. Human exposure-biokinetic models
include large numbers of parameters, which are required to describe the many processes that contribute to
Pb intake, absorption, distribution, and elimination. The large number of parameters complicates the
assessment of confidence in individual parameter values, many of which cannot be directly measured.
Statistical procedures can be used to evaluate the degree to which model outputs conform to "real-world"
observations, and values of influential parameters can be statistically estimated to achieve good
agreement with observations. Still, uncertainty can be expected to remain regarding parameters in
complex exposure-biokinetic models. Such uncertainties need to be identified and their impacts on model
predictions quantified (i.e., sensitivity analysis or probabilistic methods).
The ICRP Pb biokinetics model (Pounds and Lcggctt. 1998; Lcggctt. 1993) simulates age-
dependent kinetics of tissue distribution and excretion of Pb ingestion and inhalation intakes. This model
was originally developed for the purpose of supporting radiation dosimetry predictions, and it has been
used to develop cancer risk coefficients for internal radiation exposures to Pb and other alkaline earth
elements that have biokinetics similar to those of calcium (ICRP. 1993). Although the ICRP model has
not been validated by U.S. EPA as a regulatory model for Pb risk assessment, it has been applied in Pb
risk assessment (Abrahams et al.. 2006; Lorenzana et al.. 2005; Khourv and Diamond. 2003). and
portions of the model have been incorporated into the AALM. Although the Leggett (1993) model is
recognized as overpredicting children's blood Pb levels relative to the IEUBK model version 0.99d
(Pounds and Leggett. 1998). the Leggett model is used herein to show trends in the compartmental Pb
loadings over time in children for illustrative purposes. The AALM v2.0 (U.S. EPA. 2019b) increased the
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Leggett model rate for Pb transfer from red blood cells to plasma for the ages of 1, 5, and 10 years to
align with predicted blood Pb of children with the IEUBK model version 1.1. The mass of Pb in cortical
and trabecular bone, mass of Pb in soft tissues (kidney and liver), and total body burden of Pb were
largely unaffected by the increase in Pb transfer from red blood cells to plasma. The Leggett (1993)
model is used in this ISA to estimate compartmental Pb mass distributions rather than the AALM v2.0
since the AALM model is still under revision at this time. For figures within this appendix,
compartmental Pb concentrations were estimated using tissue and compartment masses and volumes
based on equations and parameters from O'Flaherty's studies (O'Flaherty, 1995, 1993).
The IEUBK model was designed to assess changes in blood Pb of children from birth to 7 years
of age over periods of no less than a month. Section 4.4.5 of the 2006 Pb AQCD (U.S. EPA, 2006)
introduces the IEUBK model along with components. U.S. EPA has recommended using the IEUBK
model at Superfund sites and Resource Conservation and Recovery Act Corrective Action sites to derive
a residential Preliminary Remediation Goal for Pb in soil that allows no more than a 5% probability that
children exceed a specified target BLL (U.S. EPA, 1994a). The predictive ability of the IEUBK model
has been evaluated following established guidelines for use of children's BLLs (measured in the fall to
capture peak blood Pb concentrations) paired with measurements of Pb in environmental media. The
IEUBK model v0.99 was evaluated with children's blood Pb data paired with measurements of Pb in yard
soil of their residences (Hogan et al„ 1998). The evaluation assessed the predictive ability of IEUBK to
estimate exceedances of a target BLL of 10 (ig/dL. The elimination rates of Pb from the body were
increased to the upper end of biologically plausible limits to lower IEUBK v0.99 predicted BLLs
associated with Pb in soil [see p. 1564 of White et al. (1998) and pp. 32-33 of U.S. EPA (1994b). Thus,
the rates of Pb intake from soil are aligned with elimination rates that are hard coded into the IEUBK
model. Prior to release of IEUBK v2.0, which lowered the default ingestion rates of children, the
predictive ability to estimate exceedances of target blood Pb of 5 (ig/dL (GM blood Pb 2.3 (ig/dL) was
completed using children's blood Pb data paired with concentrations of Pb in yard soil and indoor dust, as
well as bioavailability in those media (Brown et al., 2022; U.S. EPA, 2021a). SRC (2020) provides a full
description of IEUBK v2.0, including all equations and parameters.
The AALM was created to expand the capability of U.S. EPA biokinetic modeling to include
adolescents and adults and add the ability to assess the effect of intermittent Pb exposures. The AALM
uses modeling concepts taken from Leggett, O'Flaherty, and others to enhance the accuracy of the model
(U.S. EPA, 2006; Maddaloni et al„ 2005; O'Flaherty et al., 1998; O'Flaherty, 1998; Pounds and Leggett,
1998; Leggett. 1993; Leggett et al.. 1993). Section 4.4.8 of the 2006 Pb AQCD (U.S. EPA. 2006)
introduces the AALM and its components, including detailed introductions to the Leggett and O'Flaherty
models. Since that time, the AALM has gone through significant development. In September 2019, the
AALM Version 2.0 was publicly released and subsequently peer reviewed by a U.S. EPA Scientific
Advisory Board (SAB) panel. Currently, the AALM is under revision to respond to SAB review
comments. A detailed technical support document provides details related to all model parameters,
equations, and evaluations performed prior to public release of AALM v2.0 (U.S. EPA, 2019b).
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As described in Section 2.1.2, Zartarian et al. (2017) used the SHEDS-Multimedia model in
combination with an approximation of the IEUBK model to estimate drinking water Pb contributions to
blood Pb in U.S. children. In the SHEDS-IEUBK coupled methodology, the SHEDS-Multimedia model
takes the place of the exposure and variability components of the IEUBK model by generating a
probability distribution of Pb intakes across media. SHEDS-IEUBK relies on an approximation of the
IEUBK biokinetics in the form of regression equations relating Pb uptake with BLL at specified months
of life. Using these regression equations, the estimated BLL at a specified month of life assumes a
constant rate of uptake from birth to the month of interest. Thus, SHEDS-IEUBK is a slope factor model,
not a biokinetic model. When interpreting the results of IEUBK and SHEDS-IEUBK modeling, it is
important to recognize that the IEUBK model was developed, calibrated, and evaluated for site-specific
risk assessments as described above, whereas the SHEDS-IEUBK methodology uses national databases to
estimate exposure distributions for probabilistic, national-scale, population-based aggregate Pb exposure
modeling.
2.7 Summary and Conclusions
2.7.1 Exposure
Exposure information discussed in this assessment builds upon conclusions of the 2013 Pb ISA
(U.S. EPA, 2013), which built upon conclusions of the 2006 Pb AQCD (U.S. EPA, 2006). U.S.
population exposures have declined over time, as evidenced by the continued reduction in BLLs across
the United States. Sources of exposure remain, both related to air sources and other sources and pathways.
Section 2.1 details possible exposure routes for Pb. The air-related pathways for Pb exposure
occur through the inhalation of ambient air Pb, inhalation and ingestion of Pb in soil and/or resuspended
indoor or outdoor dust, ingestion of drinking water contaminated with Pb deposited from the atmosphere,
and Pb in dietary sources such as animals or plants that have taken up Pb that was deposited onto soil
with which these organisms interacted. Non-air-related pathways include exposure to corrosion
byproducts leaching into drinking water, as in the case of the Flint Water Crisis, occupational exposures
to job-related Pb, hand-to-mouth contact with consumer goods, hand-to-mouth contact with paint chips or
peeling paint containing Pb, and inhalation of dust related to Pb-containing paint or other materials in
homes and demolition of older homes containing Pb. It is difficult to ascertain the original source of Pb
involved with different exposure routes. As a result, a wide range of Pb exposures were included in this
assessment.
Environmental Pb concentrations used to estimate exposure can be collected from air monitoring,
soil Pb samples, dust Pb samples, and dietary sources including water Pb samples and food Pb samples.
Biokinetic models, such as the IEUBK and AALM models, simulate human exposure to Pb from multiple
sources and through intake routes of inhalation and ingestion. The IEUBK model was designed to assess
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changes in blood Pb of children from birth to 7 years of age over periods of no less than a month. The
predictive ability of the IEUBK model has been evaluated following established guidelines for use of
children's BLLs (measured in the fall to capture peak blood Pb concentrations) paired with measurements
of Pb in environmental media. The AALM was designed to expand the capability of biokinetic modeling
to include adolescents and adults and incorporates modeling concepts from various sources to enhance its
accuracy.
There is evidence Pb concentrations in the environment have decreased within the United States
over time. As noted in Appendix 1, air Pb concentrations in the United States have continued to decline.
The AHHS II found the number of homes with Pb-based paint has also decreased. However,
environmental Pb concentrations can vary across urban centers as a result of local meteorology, and
contributions of point/nonpoint sources of airborne Pb may still lead to exposure. Pb from resuspended
dust can also contribute to ambient air Pb concentrations. Airborne particles of Pb tend to be small but
overall have shifted to larger sizes over the past few decades as a result of the United States phasing out
leaded gasoline in automobiles. Pb particles in soil and resuspended dust tend to be coarser. Humans can
also be exposed to Pb in soil through hand-to-mouth contact. This is the main pathway of Pb exposure for
children, who play closer to the ground and in outdoor areas that may be contaminated with Pb. Finer soil
particles (<63 |im in diameter) adhere to human hands more efficiently than larger particles.
Both biological and nonbiological factors can contribute to increased exposure and associated
BLLs. EBLLs have been linked to proximity to Pb-emitting sources. Young children have higher risk of
exposure due to behaviors close to the ground and EBLLs due to higher bone turnover and lower overall
body mass. Seniors may have EBLLs because of higher lifetime Pb exposure due to higher exposures
before broad Pb regulation. Factors such as recent immigration, being a member of a racial/ethnic
minority group, and lower SES have been shown to be linked to increased BLLs. There may also be a
variety of co-contaminants present with Pb in the environment, including heavy metals, volatile
compounds, and PAHs, depending on the source and environmental pathway.
2.7.2 Toxicokinetics
The major routes of exposure to Pb are ingestion and inhalation. Both particle size and solubility
affect systemic absorption of Pb in the respiratory tract. Section 2.2.1.1 summarizes recent research on the
bioaccessibility of inhaled Pb in the lung and GI tract as a function of exposure source and particle size.
The absorption of Pb in the GI tract is influenced by physiological states of the individual and
physiochemical characteristics of the Pb-bearing material ingested. The absence of food in the GI tract
increases absorption of water-soluble Pb. Age and nutritional interactions of Pb with dietary elements,
most notably Fe, Ca2+, and Zn, may also affect GI absorption of Pb. The RBA of Pb has been tested for
different Pb forms and sources, including in swine as summarized in Table 2-10. Research has
demonstrated enrichment of Pb in smaller particle sizes (varied among studies, e.g., <50 or <100 jjxn),
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which also show greater bioaccessibility than larger particle sizes. As explored in the 2013 Pb ISA (U.S.
EPA. 2013). the majority of Pb in the body is found in bone (roughly 90% in adults, 70% in children);
only about 1% of Pb is found in blood. Pb in blood is primarily (-99%) bound to RBCs. It has been
suggested that the small fraction of Pb in plasma (<1%) may be the more biologically labile and
toxicologically active fraction of the circulating Pb. The relationship between Pb in blood and plasma is
approximately linear at relatively low daily Pb intakes and at blood Pb concentrations below -20-
30 (ig/dL and becomes curvilinear at higher blood Pb concentrations due to saturable binding to RBC
proteins. As BLL increases and the higher affinity binding sites for Pb in RBCs become saturated, a larger
fraction of the blood Pb is available in plasma to distribute to brain and other tissues.
The burden of Pb in the body may be viewed as divided between a dominant slow compartment
(bone) and smaller fast compartment(s) (soft tissues). Pb uptake and elimination in soft tissues is much
faster than in bone. Pb accumulates in bone regions undergoing the most active calcification at the time of
exposure. On the basis of Leggett (1993). trabecular bone is expected to receive 56% of Pb depositing
from plasma to bone of adults and only 20% of the Pb depositing in 1-year-olds. Cortical bone receives
44% and 80% of Pb deposited from plasma to bone in adults and 1-year-olds, respectively. A high bone
formation rate in early childhood results in the rapid uptake of circulating Pb into mineralizing bone;
however, in early childhood, bone Pb is also recycled to other tissue compartments or excreted in
accordance with a high bone resorption rate (O'Flahertv. 1995). Thus, much of the Pb acquired early in
life is not permanently fixed in the bone.
The exchange of Pb from plasma to the bone surface is a relatively rapid process. Pb in bone
becomes distributed in trabecular and dense cortical bone. The proportion of cortical to trabecular bone in
the human body varies by age but, on average, is about 80% cortical to 20% trabecular. Of the bone types,
trabecular bone is more reflective of recent exposures than cortical bone due to the slow turnover rate and
lower blood perfusion of cortical bone. Some Pb diffuses to kinetically deeper bone regions, where it is
relatively inert, particularly in adults. These bone compartments are much more labile in infants and
children than in adults, as reflected by half-times for movement of Pb from bone into to the plasma
(e.g., cortical half-time = 0.23 years at birth, 3.7 years at 15 years of age, and 23 years in adults;
trabecular half-time = 0.23 years at birth, 2.0 years at 15 years of age, and 3.8 years in adults) (Leggett.
1993).
The dominant elimination phase of Pb kinetics in the blood, exhibited shortly after a change in
exposure occurs, has a half-life of -20-30 days in adults. In children under the age of 3 years, a half-time
of only 6.4 days has been observed (Figure 2-7). An abrupt change in Pb uptake gives rise to a relatively
rapid change in blood Pb to a new quasi-steady state, achieved in -75-100 days (i.e., 3-4 times the blood
elimination half-life). A slower phase of Pb clearance from the blood may become evident with longer
observation periods following a decrease in exposure due to the gradual redistribution of Pb among bone
and other compartments.
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2.7.3
Pb Biomarkers
Trends in BLLs have been decreasing since 1976. NHANES revealed the GM blood Pb
concentration among children 1-5 years of age, based on a sample from 2017/2018, was 0.670 (ig/dL
(95% CI: 0.600, 0.748), whereas the GM blood Pb concentration among adults >20 years of age from
samples taken during the same period was 0.855 (ig/dL (95% CI: 0.816, 0.895). A GM BLL of
0.753 (ig/dL (95% CI: 0.723, 0.784) was representative of the entire U.S. population. Several studies have
shown evidence of seasonality of BLLs, with peaks occurring during summer and fall. This is attributed
to several factors, including greater soil resuspension because of drier soil conditions during the warm
season.
Because BLLs are, on average, becoming lower as a result of reductions in exposure,
methodology has had to improve to measure BLLs at lower levels of detection. At lower BLLs,
contamination of equipment can make a proportionally larger contribution to the BLL measured. Pb
contamination can occur in laboratory reagents and supplies, as well as during sample collection.
Laboratories have had to update equipment to measure at lower limits of detection from flame absorption
spectroscopy in the 1970s to newer methods, such as ICP-MS analysis used today. Capillary blood
samples are commonly collected due to their ease of collection (i.e., a finger prick) versus venipuncture
for venous blood samples. Point-of-care instruments using ASV offer low-cost, "in office" results within
minutes. However, capillary samples have been recorded as biased higher and result in more false
positives than venous blood samples. CDC recommends a venous sample should be collected if a
capillary test results in a value greater than or equal to the BLRV of 3.5 (ig/dL. Bone measurements have
advanced through the use of portable XRF, providing a less invasive way of measuring bone Pb and
spatial measurements of bone Pb to inform how Pb is incorporated into bone.
BLL is the most commonly measured Pb biomarker in literature and has been correlated to air Pb
concentrations, soil and dust Pb concentrations, and dietary Pb concentrations including tap water. BLL is
influenced by both recent and long-term exposure history, along with contributions from Pb stored in
bone. This contribution of bone Pb to blood Pb depends on duration and intensity of exposure, age, and
other physiological stressors that affect bone remodeling beyond that which normally and continuously
occurs. In children, largely due to faster exchange of Pb to and from bone, blood Pb is both an index of
recent exposure and potentially an index of body burden. In adults and children, wherein exposure to Pb
has effectively ceased or greatly decreased, a slow decline in blood Pb concentrations over a period of
years is most likely due to the gradual release of Pb from bone. Bone Pb is an index of cumulative
exposure and body burden. Even bone compartments should be recognized as reflective of differing
exposure periods, with Pb in trabecular bone exchanging more rapidly with blood than Pb in cortical
bone. This difference in the compartments makes Pb in cortical bone a better marker of cumulative
exposure and Pb in trabecular bone more likely to be correlated with blood Pb, even in adults.
The concentration of Pb in urine follows blood Pb concentration in that it mainly reflects the
exposure history of the previous few months and, therefore, is likely a relatively poor index of Pb body
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burden. There is added complexity with Pb in urine because concentration is also dependent upon urine
flow rate, which requires timed urine samples that are often not feasible in epidemiologic studies. Hair as
a biomarker has methodological issues because of contamination from environmental sources or artificial
hair treatments. The neonatal line formed in deciduous teeth during birth can be used to distinguish
between prenatal and postnatal dentine and enamel and can be used to discuss exposure history but has
not been used extensively. Other biomarkers have been used to a lesser extent (e.g., Pb in saliva).
2.7.4 Air Pb-Blood Pb Relationships
Table 2-13 provides summaries of studies that measured air-to-blood Pb slopes in children. There
is variability in study location, population, air and blood Pb concentrations, and analysis used among
studies. Studies have described the blood Pb-air Pb slope as either log-log (Meng et al.. 2014; Richmond-
Bryant et al.. 2013; Zahran et al.. 2013a; Bierkens et al.. 2011; Schnaas et al.. 2004; Haves et al.. 1994;
Brunekreef. 1984) or linear (Hilts. 2003; Tripathi et al.. 2001; Schwartz and Pitcher. 1989). Much of the
earlier literature on slope factors was summarized by Brunekreef (1984). who performed a meta-analysis
using many of the relevant references in the 1986 AQCD (U.S. EPA. 1986) and found blood Pb versus air
Pb slope |3 was smaller at high blood and air levels.
Newer studies after the time of leaded gasoline usage and not focused on communities near
significant air Pb sources show increasing slope factors with decreasing air Pb concentrations.
Figure 2-16 shows a range of slope factors as a function of air concentration data sets, including those of
recent studies. Richmond-Bryant et al. (2014) compared NHANES regression results with those from the
literature and found the slope factor increases with decreasing air Pb among children 0-11 years of age.
Using 1999-2008 NHANES BLL data, Meng et al. (2014) found BLL was more consistently and strongly
associated with PMio than either TSP or PM2.5, which had an appreciably smaller number of samples.
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_ EPA/600/R-23/375
£% United States
Environmental Protection January 2024
m m Agency www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 3: Nervous System Effects
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
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CONTENTS
DOCUMENT GUIDE 3-iii
LIST OF TABLES 3-v
LIST OF FIGURES 3-vii
ACRONYMS AND ABBREVIATIONS 3-viii
APPENDIX 3 NERVOUS SYSTEM EFFECTS 3-1
3.1 Introduction 3-2
3.2 Scope 3-2
3.3 Biological Plausibility 3-4
3.4 Overt Nervous System Toxicity 3-14
3.4.1 Epidemiologic Studies of Brain Structure and Function 3-14
3.4.2 Experimental Animal Studies of Brain Structure and Function 3-18
3.4.3 Integrated Summary of Overt Nervous System Toxicity 3-25
3.5 Nervous System Effects Ascertained during Childhood, Adolescent, and Young Adult
Lifestages 3-26
3.5.1 Cognitive Function in Children 3-26
3.5.2 Externalizing Behaviors: Attention, Impulsivity, and Hyperactivity in Children 3-87
3.5.3 Externalizing Behaviors: Conduct Disorders, Aggression, and Criminal Behavior in
Children, Adolescents, and Young Adults 3-114
3.5.4 Internalizing Behaviors: Anxiety and Depression in Children 3-126
3.5.5 Motor Function in Children 3-139
3.5.6 Sensory Organ Function in Children 3-153
3.5.7 Social Cognition and Behavior in Children 3-162
3.6 Nervous System Effects Ascertained during Adult Lifestages 3-175
3.6.1 Cognitive Function in Adults 3-175
3.6.2 Psychopathological Effects in Adults 3-194
3.6.3 Sensory Organ Function in Adults 3-204
3.6.4 Neurodegenerative Diseases 3-212
3.7 Evidence Inventories - Data Tables to Summarize Study Details 3-228
3.8 References 3-502
3-iv
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LIST OF TABLES
Table 3-1 Statistics associated with the international pooled analysis of data from seven cohort
studies
3-29
Table 3-2 Summary of evidence Indicating a causal relationship between Pb exposure and cognitive
effects in children 3-84
Table 3-3
Table 3-4
Table 3-5
Table 3-6
Summary of evidence indicating a causal relationship of Pb exposure with attention,
impulsivity, and hyperactivity 3-110
Summary of evidence for a likely to be causal association between Pb exposure and
conduct disorders, aggression, and criminal behavior in children and adolescents
Summary of evidence for a likely to be causal relationship between Pb exposure and
internalizing behaviors in children
3-124
3-137
Summary of evidence indicating a likely to be causal relationship between Pb exposure
and motor function in children 3-151
Table 3-7 Evidence that is suggestive of, but not sufficient to infer, a causal relationship between Pb
exposure and sensory organ function in children 3-160
Table 3-8 Evidence that is suggestive of, but not sufficient to infer, a causal relationship between Pb
exposure and social cognition and behavior in children 3-173
Table 3-9 Summary of evidence for a causal relationship between Pb exposure and cognitive effects
in adults 3-192
Table 3-10
Table 3-11
Table 3-12
Table 3-1E
Table 3-1T
Table 3-2E
Table 3-3E
Table 3-4E
Table 3-4T
Table 3-5E
Table 3-6E
Summary of evidence for a likely to be causal relationship between Pb exposure and
psychopathological effects in adults
Summary of the evidence that is suggestive of, but not sufficient to infer, a causal
relationship between sensory function in adults
Epidemiologic studies of Pb exposure and overt nervous system toxicity
Animal toxicological studies of Pb exposure and brain function
Epidemiologic studies of Pb exposure and full-scale intelligence quotient_
Epidemiologic studies of Pb exposure and infant development
Animal toxicological studies of Pb exposure and cognitive function _
Epidemiologic studies of Pb exposure, academic performance, and achievement
3-203
3-210
Summary of evidence that is suggestive of, but not sufficient to infer, a causal relationship
between Pb exposure and neurodegenerative diseases in adults 3-226
3-228
3-233
3-258
3-276
Epidemiologic studies of Pb exposure and performance on neuropsychological tests of
cognitive function, i.e., learning, memory, and executive function 3-283
3-293
3-321
Epidemiologic studies of Pb exposure and cognitive effects: population or group mean
blood Pb levels >5 [jg/dL 3-329
3-v
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Table 3-7E Epidemiologic studies of Pb exposure and performance on neuropsychological tests of
attention, impulsivity, and hyperactivity, ADHD-related behaviors, and clinical ADHD in
children 3-347
Table 3-7T Animal toxicological studies of Pb exposure and externalizing and internalizing behaviors 3-365
Table 3-8E Epidemiologic studies of Pb exposure and performance on neuropsychological tests of
attention, impulsivity, and hyperactivity, attention deficit/hyperactivity disorder-related
behaviors, and clinical attention deficit/hyperactivity disorder in children; group or
population mean blood Pb level >5 [jg/dL, any study design 3-373
Table 3-9E Epidemiologic studies of Pb exposure and externalizing behaviors including conduct
disorders, aggression, and criminal behavior in children and adolescents 3-384
Table 3-10E Epidemiologic studies of Pb exposure and internalizing behaviors in children 3-398
Table 3-11E Epidemiologic studies of Pb exposure and motor function in children 3-408
Table 3-11T Animal toxicological studies of Pb exposure and motor function 3-424
Table 3-12E Epidemiologic studies of Pb exposure and sensory organ function in children 3-432
Table 3-13E Epidemiologic studies of Pb exposure, social cognition, and behavior in children 3-441
Table 3-14E Epidemiologic studies of exposure to Pb and cognitive function in adults 3-451
Table 3-15E Epidemiologic studies of Pb exposure and psychopathological effects in adults 3-461
Table 3-16E Epidemiologic studies of Pb exposure and sensory organ function in adults 3-468
Table 3-16T Animal toxicological studies of Pb exposure and sensory organ function 3-476
Table 3-17E Epidemiologic studies of exposure to Pb and neurodegenerative disease in adults 3-478
Table 3-17T Animal toxicological studies of Pb exposure and neurodegeneration 3-498
3-vi
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LIST OF FIGURES
Figure 3-1 Potential biological pathways for nervous system effects following developmental
exposure to Pb. 3-5
Figure 3-2 Potential biological pathways for nervous system effects following postweaning exposure
to Pb. 3-6
Figure 3-3 The relationship between blood Pb level at age 11 and brain outcomes in adulthood. 3-16
Figure 3-4 Associations between blood Pb levels and full-scale intelligence quotient in children. 3-32
Figure 3-5 Associations between biomarkers of Pb exposure and Bayley Score of Infant
Development Mental Development Index. 3-39
Figure 3-6 Association of blood Pb level with reading and math scores among North Carolina school
children (average across all grades). Left panel displays impact of blood Pb level on math
test score. Right panel displays impact of blood Pb level on reading test score. 3-61
Figure 3-7 Relationship between concurrent blood Pb level and intelligence quotient among Italian
adolescents using a cubic spline fit. 3-63
Figure 3-8 Relationship between log-transformed blood Pb level and intelligence quotient using an
ordinary least squares fit. 3-64
Figure 3-9 Two distributions of intelligence test scores demonstrating the consequence in a small
shift in the mean score. 3-71
Figure 3-10 Scatter plots and regression lines of blood Pb level and 18-month Mental Developmental
Index among children in manganese (A) quintiles 1-4 and (B) quintile 5. 3-75
Figure 3-11 Mean ± standard deviation behavior performance in the Go/No-Go task according to
quartiles of exposure for (A and B) cord blood Pb and (C) childhood blood Pb level at age
11 years. 3-93
Figure 3-12 Associations of monthly airborne Pb exposure levels from birth to age 12 with scores for
anxiety and depression behaviors on the Behavior Assessment System for Children. 3-129
Figure 3-13 Associations between biomarkers of Pb exposure and Bayley Score of Infant
Development Psychomotor Developmental Index. 3-141
Figure 3-14 Differences in mean difference tooth Pb levels for autism spectrum disorder in discordant
twin pairs versus (A) non-autism spectrum disorder twin pairs or (B) autism spectrum
disorder concordant twin pairs. 3-163
Figure 3-15 Hazard rate ratios for Alzheimer's disease mortality by blood Pb level including the lower
95% confidence interval. 3-222
3-vii
-------
ACRONYMS AND ABBREVIATIONS
AA atomic absorption BPAQ
AAS atomic absorption spectrometry BrainAGE
Ap amyloid beta BRIEF
ABR auditory brainstem response
AD Alzheimer's disease BRIEF-A
ADD attention deficit disorder
BRIEF-P
ADHD attention deficit/hyperactivity disorder
ADHD-RS ADHD rating scale
ADOS Autism Diagnostic Observation BRS
Schedule ggj
ADRA2A alpha-2A-adrenergic receptor BSID
ALAD aminolevulinic acid dehydratase
ALS Amyotrophic Lateral Sclerosis BSID-IIS
ALSPAC Avon Longitudinal Study of Parents
and Children BT20+
AOR adjusted odds ratio Ca2+
APP amyloid precursor protein CANTAB
AQCD Air Quality Criteria Document
As arsenic CAR
avg average CARES
ASD autism spectrum disorder
CARS
ASQ:I Ages and Stages Questionnaire
Inventory CAT
ASSQ Autism Spectrum Screening CBCL
Questionnaire CBLI
ATP adenosine triphosphate CCAAPS
BAARS Barkley Adult ADHD-IV Rating Scale
BACE1 beta-secretase 1 CCEI
BAEP brainstem auditory evoked potential Cd
BASC Behavior Assessment System for CDIIT
Children
BASC-2 Behavior Assessment System for CDK5
Children, second revision Ce
BBB blood-brain barrier CEM
BDI Beck Depression Inventory CERAD
BDNF brain-derived neurotrophic factor
BKMR Bayesian kernel machine regression CHECK
BKT Binet Kamat T est
BLL blood lead level CHEER
BMD benchmark dose ^TT, r„
CHMS
BMDL benchmark dose lower 95% confidence
limit U
BMI body mass index
BMS Baltimore Memory Study
CKJJ
BNT Boston Naming Test
CKiD
BPA bisphenol A
Buss-Perry Aggression Questionnaire
Brain Age Gap Estimation
Behavior Rating Inventory of
Executive Functions
Behavior Rating Inventory of
Executive Functions for Adults
Behavior Rating Inventory of
Executive Functions for Preschool
Children
behavioral rating scale
Behavioral Symptoms Index
Bay ley Scales of Infant and Toddler
Development
Bay ley Scales of Infant and Toddler
Development - Spanish Version
Birth to Twenty Plus
calcium ion(s)
Cambridge Neuropsychological Test
Automated Battery
Cortisol awakening response
Communities Actively Researching
Exposure Study
Childhood Autism Rating Scale
catalase
Child Behavior Check List
cumulative blood lead index
Cincinnati Childhood Allergy and Air
Pollution Study
Crown-Crisp Experiential Index
cadmium
Comprehensive Developmental
Inventory for Infants and Toddlers
cyclin-dependent kinase 5
cesium
Coarsened Exact Matching
Consortium to Establish a Registry for
Alzheimer's Disease
Children's Health and Environmental
Chemicals in Korea
Children's Health and Environmental
Research
Child Health Monitoring System
confidence interval
Composite International Diagnostic
Interview
chronic kidney disease
Chronic Kidney Disease in Children
Cincinnati Lead Study
3-viii
-------
CNS central nervous system
Co cobalt
C-P central-to-peripheral
cpd cycles per degree
CPR Conditioned Position Responding
CPRS Conners' Parent Rating Scale
CPRS-R Conners' Parent Rating Scale-Revised
CPT Continuous Performance Test
C-R concentration-response
CR chromium
CREB cyclic adenosine 3',4'-monophosphate
response element binding protein
CRISYS-R Crisis in Family Systems-Revised
CRP C-reactive protein
CRS Conners' Rating Scale
CRS-R Conners' Rating Scale-Revised
CRT Combined Raven's Test
CSF cerebrospinal fluid
C-TRF Caregiver-T eacher Report F orm
CTRS Conners'Teacher Rating Scale
CTRS-R Conners'Teacher Rating Scale-
Revised
C-V R2 cross validated R-square
CVA cerebrovascular accident
CVD cardiovascular disease
CVLT California Verbal Learning Test
CVLT-C California Verbal Learning Test-
Children's Version
d day(s)
DAT1 dopamine transporter
DBD Disruptive Behavior Disorder
DDE dichlorodiphenyldichloroethylene
DI deionized
DISCI Disrupted-in-Schizophrenia-1
DMTS Delayed Matching-to-Sample
DQ development quotient
DRD2 Dopamine Receptor D2
DNAm DNA methylation
DSC Digit Symbol Coding
DSM Diagnostic and Statistical Manual of
Mental Disorders
DSST Digit Symbol Substitution Test
DTI Diffusion Tensor Imaging
ECAT elemental carbon attributable to traffic
ECDI Early Child Development Inventory
EE effect estimate
EEG electroencephalogram
ELEMENT Early Life Exposure in Mexico to
Environmental Toxicants
EMOCI emotional regulation
EOG end of grade
EPM elevated plus maze
EPN early postnatal
EPSC excitatory postsynaptic currents
ERG electroretinography
ERP event-related potential
ETS environmental tobacco smoke
F female
F# filial generation
FA fractional anisotropy
FBB-ADHS Fremdbeurteilungsbogen fur
Aufmerksamkeitsdefizit/Hyperaktivitat
storungen
Fe iron
FFQ Food Frequency Questionnaire
FI fixed interval
FLEHS Flemish Environment and Health Study
FR fixed ratio
FSIQ full-scale intelligence quotient
FST forced swim test
GABA gamma-aminobutyric acid
GCNT1 glucosaminyl (N-acetyl) transferase 1
GD gestational day
GDS Gesell Developmental Schedules
GFAAS graphite furnace atomic absorption
spectrometry
GMR geometric mean ratio
GRIN glutamate ionotropic receptor N-methyl
D-aspartate-type subunit
GSH glutathione
GSI Global Severity Index
GST glutathione S-transferase
HCB hexachlorobenzene
HDL high-density lipoprotein
HFE hemochromatosis gene
Hg mercury
Hgb hemoglobin
HHANES Hispanic Health and Nutrition
Examination Survey
HI hyperactivity and impulsivity
HNES Home Nurture Environment Scale
HNRS Heinz Nixdorf Recall Study
HOME Health Outcomes and Measures of the
Environment
HPA hypothalamic pituitary adrenal
hr hour(s)
HR hazard ratio
HR-ICP-MS high resolution inductively coupled
plasma mass spectrometry
3-ix
-------
HRR hazard rate ratio
HRT hormone replacement therapy
ICD International Classification of Diseases
ICP-DRC-MS dynamic reaction cell for inductively
coupled plasma mass spectrometry
ICP-MS inductively coupled plasma mass
spectrometry
ICP-OES inductively coupled plasma optical
emission spectroscopy
ICP-SFMS inductively coupled plasma sector field
mass spectrometry
INMA INfancia y Medio Ambiente
IQ intelligence quotient
IQR interquartile range
ISA Integrated Science Assessment
ISAT Illinois Standard Achievement Test
K6 Kessler Psychological Distress Scale
K-ABC Kaufman Assessment Battery for
Children
K-ARS Korean ADHD Rating Scale
K-CBCL Korean Child Behavior Check List
KEDI Korean Educational Development
Institute
KiTAP Test of Attentional Performance for
Children
KNHANES Korea National Health and Nutrition
Examination Survey
K-SADS Kiddie Schedule for Affective
Disorders and Schizophrenia
K-SADS-PL-K Kiddie Schedule for Affective
Disorders and Schizophrenia Present
and Lifetime - Korean Version
K-XRF K-shell X-ray fluorescence
LASSO least absolute shrinkage and selection
operator
In natural log
LOD limit of detection
LTP long-term potentiation
LURF Land Use Random Forest
M male
Mat maternal
MAT Metropolitan Achievement Test
MCU mitochondrial Ca2+ uniporter
MDAT Malawi Development Assessment Tool
MDI Mental Development Index
mDISCl mouse Disrupted-in-Schizophrenia-1
ME maternal exposure
MEAP Michigan Educational Assessment
Program
MeHg methyl mercury
MHI-5 Mental Health Index 5-item
MIREC Maternal-Infant Research on
Environmental Chemicals
MMSE Mini Mental State Examination
Mn manganese
mo month(s)
MOCEH Mothers' and Children's
Environmental Health
MRI magnetic resonance imaging
MrOS Osteoporotic Fractures in Men Study
MRS magnetic resonance spectroscopy
MSCA McCarthy Scales of Children's
Abilities
NaAc sodium acetate
NAS Normative Aging Study
NBAS Neonatal Behavioral Assessment
Scales
NBNA Neonatal Behavioral Neurological
Assessment
NCDS Nunavik Child Development Study
NEI National Emissions Inventory
NHANES National Health and Nutrition
Examination Survey
NHBCS New Hampshire Birth Cohort Study
NHS Nurses' Health Study
NMDAR N-methyl-D-aspartate receptor
NPR Norwegian Patient Registry
NR not reported
NS no stress
OD/CD oppositional defiant and conduct
disorder
OFT open-field test
OLS ordinary least squares
OR odds ratio
ORIEN orientation/engagement
OTB operant test battery
Pb lead
PbO lead oxide
PC primary caregiver
PCBs polychlorinated biphenyls
PCNA proliferating cell nuclear antigen
PD Parkinson's disease
PDI Psychomotor Developmental Index
PECOS Population, Exposure, Comparison,
Outcome, and Study Design
PEG Parkinson's Environment and Genes
PERI perinatal
PHDCN Project on Human Development in
Chicago Neighborhoods
PIQ Performance Intelligence Quotient
PIR poverty-income ratio
3-x
-------
PM2.5 fine particulate matter
PND postnatal day
PPI Psychopathic Personality Inventory
PR prevalence ratio
PROGRESS Programming Research in Obesity,
Growth, Environment and Social
Stressors
PRP post-reinforcement pause
PTA pure-tone average
pts points
p-tau phosphorylated tau
PTSD post-traumatic stress disorder
PW postweaning
Q quartile
RNS reactive nitrogen species
RO DI reverse osmosis deionized
ROS reactive oxygen species
RR relative risk
RSEI Risk Screening Environmental
Indicators
SCN suprachiasmatic nucleus
SCWT Stroop Color-Word Test
SD standard deviation
SDQ Strengths and Difficulties
Questionnaire
Se selenium
SE standard error
SES socioeconomic status
SGA small for gestational age
SGPD System Genomics of Parkinson's
Disease
SMBCS Sheyang Mini Birth Cohort Study
SMS Social Maturity Scale
SOD superoxide dismutase
Sp specificity protein
SPHERL Study for Promotion of Health in
Recycling Lead
SPM Standard Progressive Matrix
SQ social quotient
SRP self-report of personality
SRS Social Responsiveness Scale
SWAN Strengths and Weaknesses of ADHD
Symptoms and Normal Behavior Scale
T# trimester #
TBD to be determined
TBPS Taiwan Birth Panel Study
TEACh Test of Everyday Attention for
Children
TMT Trail Making Test
TOKS tin-ore kilns and smelters
TRD Temporal-Response Differentiation
TRF Teacher Report Form
TSCD Tohoku Study of Child Development
TST Tail Suspension Test
TUNEL terminal deoxynucleotidyl transferase
dUTP nick end labeling
U.S. EPA United States Environmental Protection
Agency
USV ultrasonic vocalizations
VA visual acuity
VDR vitamin D receptor
VEP visual evoked potential
VIF variance inflation factor
VIQ Verbal Intelligence Quotient
VMI visual-motor integration
WAIS Weschler Adult Intelligence Scale
WASI Wechsler Abbreviated Scale of
Intelligence
wk week(s)
WHO World Health Organization
WIAT Wechsler Individual Achievement Test
WISC Wechsler Intelligence Scale for
Children
WJTA Woodcock-Johnson Test of
Achievement
WMC working memory capacity
WMH white matter hyperintensities
WMS Weschler Memory Scale
WPPSI Wechsler Preschool and Primary Scale
of Intelligence
WRAML Wide Range Assessment of Memory
and Learning
WRAT Wide Range Achievement Test
XRF X-ray fluorescence
yr year(s)
YSR youth self-report
Zn zinc
3-xi
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APPENDIX 3 NERVOUS SYSTEM EFFECTS
Summary of Causality Determinations for Pb Exposure and Nervous System Effects
This appendix characterizes the scientific evidence that supports causality determinations for
lead (Pb) exposure and nervous system effects. The types of studies evaluated within this appendix are
consistent with the overall scope of the ISA as detailed in the Process Appendix (see Section 12.4). In
assessing the overall evidence, the strengths and limitations of individual studies were evaluated based
on scientific considerations detailed in Table 12-5 of the Process Appendix (Section 12.6.1). More
details on the causal framework used to reach these conclusions are included in the Preamble to the ISA
(U.S. EPA. 2015). The evidence presented throughout this appendix supports the follow ing causalitv
conclusions:
Outcome Group
Causality Determination
Nervous System Effects Ascertained during Childhood, Adolescent, and Young Adult Lifestages
Cognitive Effects
Causal
Attention, Impulsivity and Hyperactivity
Causal
Conduct Disorders, Aggression, and Criminal
Behavior
Likely to be causal
Internalizing Behaviors
Likely to be causal
Motor Function
Likely to be causal
Sensory Function
Suggestive of, but not sufficient to infer, a causal
relationship
Social Cognition and Behavior
Suggestive of, but not sufficient to infer, a causal
relationship
Nervous System Effects Ascertained during Adult Lifestages
Cognitive Effects
Causal
Psychopathological Effects
Likely to be causal
Sensory Function
Suggestive of, but not sufficient to infer, a causal
relationship
Neurodegenerative Disease
Suggestive of, but not sufficient to infer, a causal
relationship
The Executive Summary, Integrated Synthesis, and all other appendices of this Pb ISA can be found at
https://assessments.epa.aov/isa/document/&deid=359536
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3.1 Introduction
While Pb affects nearly every organ system, the nervous system appears to be one of the most
sensitive targets. The sections that follow provide an evaluation of the most policy-relevant scientific
evidence relating to the effects of lead (Pb) exposure on the nervous system. To maximize transparency
regarding the studies included in the appendix, the scope is defined in Section 3.2. Section 3.3, Biological
Plausibility, provides an overview of the biological pathways that potentially underlie the nervous system
effects discussed in subsequent sections of the appendix. Section 3.4 summarizes overt nervous system
toxicity, including changes in brain structure and function. There is no causality determination in this
section; rather, data presented in the section may be referenced in the outcome-specific "Summary and
Causality Determination" discussions in later sections if they provide support for the conclusions.
Sections 3.5 and 3.6 describe the epidemiologic and experimental animal evidence that pertains to
specific endpoints or outcome groupings, which are organized by the lifestage at which they are
ascertained (i.e., childhood, adolescence, and young adult [Section 3.5] and adult [Section 3.6] lifestages).
The strongest and most policy-relevant evidence within each section is discussed first. Within
Section 3.5, which focuses on exposures and outcomes ascertained during childhood lifestages, including
adolescence and early adulthood, the strongest evidence that is best substantiated at the lowest exposure
levels relates to Cognitive Effects (Section 3.5.1) and Attention, Impulsivity, and Hyperactivity (Section
3.5.2) in children. Conduct Disorders are discussed in Section 3.5.3, followed by Anxiety and Depression
(Section 3.5.4), Motor Function (Section 3.5.5), Sensory Organ Function (Section 3.5.6), and Social
Cognition and Behavior (Section 3.5.7). The next section (Section 3.6) includes endpoints that are
ascertained during adult lifestages. The section begins with an assessment of the evidence pertaining to
Cognitive Effects in Adults (Section 3.6.1) followed by sections on Anxiety, Depression, and
Psychopathological Effects (Section 3.6.2), Sensory Function (Section 3.6.3), and Neurodegenerative
Diseases (Section 3.6.4). Within each section, the collective body of evidence is integrated within and
across scientific disciplines, and issues relevant for interpreting the scientific evidence as well as the
rationale for the causality determination are outlined for relevant endpoints or outcome groupings.
3.2 Scope
The scope of this appendix is defined by Population, Exposure, Comparison, Outcome, and Study
Design (PECOS) statements. The PECOS statements define the objectives of the review and establish
study inclusion criteria, thereby facilitating identification of the most relevant literature to inform the
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Lead Integrated Science Assessment (Pb ISA).1 In order to identify the most relevant literature, the body
of evidence from the 2013 Pb ISA was considered in the development of the PECOS statements for this
appendix. Specifically, well-established areas of research; gaps in the literature; and inherent uncertainties
in specific populations, exposure metrics, comparison groups, and study designs identified in the 2013 Pb
ISA inform the scope of this appendix. The 2013 Pb ISA used different inclusion criteria than the current
ISA, and the studies referenced therein often do not meet the current PECOS criteria (e.g., due to higher
or unreported biomarker levels). Studies that were included in the 2013 Pb ISA, including many that do
not meet the current PECOS criteria, are discussed in this appendix to establish the state of the evidence
prior to this assessment. Except for supporting evidence used to demonstrate the biological plausibility of
Pb-associated nervous system effects, recent studies evaluated and subsequently discussed within this
appendix were only included if they satisfied all components of the following discipline-specific PECOS
statements:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect;
Exposure: Exposure to Pb2 as indicated by biological measurements of Pb in the body, with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure,3 or intervention groups in randomized trials and quasi-experimental studies;
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles);
Outcome: Nervous system effects including but not limited to cognitive function (e.g.,
intelligence quotient [IQ] decrement), externalizing and internalizing behaviors,
psychopathological effects, sensory organ function, motor function, and neurodegenerative
diseases; and
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
'The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articfes (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
2Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area that was of
particular relevance to the National Ambient Air Quality Standards review (e.g., longitudinal studies designed to
examine recent versus historical Pb exposure).
3Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less man or equal to 10 |im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 (im3 (PM2.5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is
difficult to assess the representativeness of these concentrations to population exposure (Section 2.5.3 (U.S. EPA.
2013)). Moreover, data illustrating the relationships of Pb-PMio and Pb-PNfc.s with blood Pb levels (BLLs) are
lacking.
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endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages);
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a BLL of 30 (ig/dL or below;4,5
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control;
Outcome: Nervous System effects; and
Study Design: Controlled exposure studies of animals in vivo.
3.3 Biological Plausibility
This section describes biological pathways that potentially underlie nervous system effects
resulting from exposure to Pb. Timing of exposure is important for the health effects for Pb. Exposures
during development can lead to improper formation and maturation of the nervous system and exposures
to the mature nervous system can lead to neurodegeneration. Figure 3-1 and Figure 3-2 graphically depict
these proposed pathways for health effects resulting from developmental exposure to Pb and later life
exposures, respectively. Proposed pathways are presented as a continuum of responses, connected by
arrows, which may ultimately lead to the apical nervous system health effects associated with exposures
to Pb at concentrations observed in epidemiologic studies. This discussion of "how" exposure to Pb may
lead to effects on the nervous system contributes to an understanding of the biological plausibility of
epidemiologic results evaluated throughout this appendix. Most of the studies cited in this subsection are
discussed in greater detail elsewhere in this appendix. The biological plausibility for Pb-induced effects
on the nervous system is supported by evidence from the 2013 Pb ISA and by recent evidence. Note that
the structure of the biological plausibility sections and the role of biological plausibility in contributing to
the weight-of-evidence analysis used in the current ISA are discussed in Section IS.7.2.
4Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
5This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLL.
The 95th percentile of the 2011-2016 National Health and Nutrition Examination Survey distribution of BLL in
children (1-5 years; n = 2,321) is 2.66 (ig/dL (Eganet al.. 2021) and the proportion of individuals with BLL that
exceed this concentration varies depending on factors including (but not limited to) housing age, geographic region,
and a child's age, sex, and nutritional status.
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Altered
Transcriptional
Regulation
Altered
Neurodevelopmental
Processes
Impaired Blood
Brain & Blood
Cerebrospinal
Fluid Barrier
Permeability
^ Neurodevelopmental
Disorders
Impaired Behavior/
Cognitive/ Function
Mood Disorders
Altered
Neurotransmitter
Signaling
Protein Binding
Pb Exposure
Impaired sensory
function
Neurodegenerative
Diseases
Note: The boxes above represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and the arrows indicate a proposed relationship
between those effects. Solid arrows denote evidence of essentiality as provided, for example, by an inhibitor of the pathway used in an experimental study involving Pb exposure.
Dotted arrows denote a possible relationship between effects. Shading around multiple boxes is used to denote a grouping of these effects. Arrows may connect individual boxes,
groupings of boxes, and individual boxes within groupings of boxes. Progression of effects is generally depicted from left to right and color coded (white, exposure; green, initial effect;
blue, intermediate effect; orange, effect at the population level or a key clinical effect). Here, population-level effects generally reflect results of epidemiologic studies. When there are
gaps in the evidence, there are complementary gaps in the figure and the accompanying text below. The structure of the biological plausibility sections and the role of biological
plausibility in contributing to the weight-of-evidence analysis used in the 2022 Pb ISA are discussed in Section IS.7.2. Source: (Shadbeciian et ai., 2019).
Figure 3-1 Potential biological pathways for nervous system effects following developmental exposure to
Pb,
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Altered Neurotransmitter
Signaling
Altered Calcium Signaling
Impaired Blood
Brain & Blood
Cerebrospinal
Fluid Barrier
Permeability
Note: The boxes above represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and the arrows indicate a proposed relationship
between those effects. Solid arrows denote evidence of essentiality as provided, for example, by an inhibitor of the pathway used in an experimental study involving Pb exposure.
Dotted arrows denote a possible relationship between effects. Shading around multiple boxes is used to denote a grouping of these effects. Arrows may connect individual boxes,
groupings of boxes, and individual boxes within groupings of boxes. Progression of effects is generally depicted from left to right and color coded (white, exposure; green, initial effect;
blue, intermediate effect; orange, effect at the population level or a key clinical effect). Here, population-level effects generally reflect results of epidemiologic studies. When there are
gaps in the evidence, there are complementary gaps in the figure and the accompanying text below. The structure of the biological plausibility sections and the role of biological
plausibility in contributing to the weight-of-evidence analysis used in the 2022 Pb ISA are discussed in Section I.S.7.2. Source: CShadbegian et ai.. 20191.
Figure 3-2 Potential biological pathways for nervous system effects following postweaning exposure to Pb.
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Plausible pathways connecting Pb exposure to apical events resulting from developmental and
later life exposures to Pb are proposed in Figure 3-1 and Figure 3-2, respectively. The proposed pathways
supported by the strongest evidence include the direct actions of Pb on cellular protein function and
subsequent initiation of oxidative stress-mediated pathways.
When Pb accumulates in the CNS, it can interfere with coordination of metal ions, which is
essential for the structure and function of many cellular proteins. Pb ions can compete with and displace
physiologically relevant ions (including Fe, Zn, Ca, and others) within proteins, leading to both altered
protein structure and function. As described in the 2013 Pb ISA, there is evidence that this ionic mimicry
and imbalance occurs in multiple organ systems, including the brain, and in proteins that perform diverse
functions including metabolism, inflammation, and oxidative stress responses. For example, Pb treatment
can disrupt Ca2+ signaling through interactions with calmodulin, voltage-gated Ca2+ channels, and various
adenosine triphosphate (ATP)ases (U.S. EPA, 2013). There is also evidence that Pb can replace Zn ions
in Zn finger-binding motifs, which are present in several transcription regulating proteins (U.S. EPA,
2013). Some research supports an interactive effect between Fe status and Pb exposure due to shared
metabolic and physiological profiles. Lifetime exposure to Pb in rats has been shown to affect Fe status
by increasing Fe content in the cortex and hippocampus of adult and aged animals and altering the
expression of divalent metal transporters (such as divalent metal transporter 1 and ferroportin) in the brain
(Zhu et al., 2013), suggesting that Pb may interfere with Fe trafficking in the brain. Often the effect of Pb
can be reduced with exogenous supplementation of biologically relevant metals. Recent studies support
the protective role of supplementation of Ca2+ (Basha and Reddy, 2015; Gottipolu and Davuljigari, 2014),
Zn (Pedroso et al., 2017), Fe (Liu et al„ 2013c) or essential metal mixtures (Basha et al„ 2014) on
neurologic alterations from Pb. These data support the hypothesis that direct competition of Pb with
metals can cause neurologic effects.
The brain has the highest energy demand and metabolism of any organ. Because of this fact,
energy homeostasis is critical and energy imbalance can increase the brain's susceptibility to stressors and
cell death. Pb-induced alterations in energy production and metabolism have been measured in several
ways. As discussed in the 2013 Pb ISA, Pb exposure can alter many aspects of energy metabolism, with
animal models demonstrating effects following both developmental and adult exposures to Pb (discussed
in Section 3.4.2.1). In recent studies of developmental Pb exposure, Pb-induced impairments in energy
production throughout the body have been measured as reductions in the activity of glucose and glycogen
metabolizing enzymes (Baranowska-Bosiacka et al., 2017) and alterations in the number and structure of
mitochondria (Ouvang et al., 2019; Gassowska et al., 2016a). Studies of Pb exposure in postweaning
animals showed similar reductions of metabolizing enzyme activity (Yun et al„ 2019; Verma et al„ 2005;
Yun and Hover, 2000; Sterling et al„ 1982), altered mitochondrial structure (Ouvang et al., 2019;
Dabrowska et al„ 2015; Sun et al., 2014), and ATPase activity (Thangarajan et al., 2018), suggesting
alteration of energy metabolism may occur regardless of the timing of Pb exposure.
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Energy production involves the formation of reactive intermediate species including reactive
oxygen (ROS) and nitrogen species (RNS). Disruptions in the mitochondria and energy metabolism result
in increased levels of ROS and RNS. While ROS are a part of normal cellular functioning, uncontrolled
production or reduced elimination of ROS by antioxidant systems can result in oxidative stress and
cellular damage (for example, DNA damage, oxidization of cellular components). Evidence reviewed in
the 2013 Pb ISA suggests that Pb may exert toxicity by disrupting cellular metabolism, increasing
ROS/RNS concentrations, and depleting antioxidant capacity (U.S. EPA, 2013). Numerous recent studies
have reported dysregulation of oxidative stress concurrent with altered mitochondrial function
(Ahmad et al., 2020; Karri et al., 2018; Maiti et al„ 2017; Kumar and Muralidhara, 2014; Baranowska-
Bosiacka et al., 2011), adding to the body of evidence. A study by Yang et al. (2014) found that Pb
downregulated the mitochondrial Ca2+ uniporter (MCU), resulting in increased ROS production in both
SH-SY5Y cells and in newborn rats. Yang and colleagues found that in vitro activation or overexpression
of MCU prevented Pb-induced oxidative stress whereas MCU inhibition or knockdown potentiated the
effects suggesting that alterations in mitochondrial function were responsible for Pb-induced ROS
production. In a similar manner, Pb has been shown to upregulate cyclophilin D, a protein that regulates
mitochondrial membrane potential, and in vitro knockdown or inhibition of cyclophilin D prevents the
Pb-induced loss of mitochondrial membrane potential (Ye et al., 2020; Ye et al., 2016a). Mitochondrial
function is also thought to be dependent on a dynamic balance between mitochondrial fission and fusion.
In a recent study, Pb reduced energy production and respiration while increasing mitochondrial ROS and
altering the expression of genes involved with mitochondrial dynamics both in vitro and in vivo
(Dabrowska et al„ 2015). In this study, knockdown of the transcription factor peroxisome proliferator-
activated receptor-y coactivator la, which protects the mitochondrial fusion and fission balance, increased
in vitro ROS production in response to Pb, further suggesting that altered mitochondrial activity results in
ROS production in response to Pb. Together, these data provide evidence that mitochondrial dysfunction
and altered energy metabolism is a source of oxidative stress.
Given their reactive nature, ROS and RNS can damage cellular proteins, lipids, and nucleic acids,
which can lead to functional and downstream signaling impairment. As discussed in the 2013 Pb ISA, Pb
exposure leads to elevated levels of ROS in neurons and other brain cells of exposed animals. Levels of
oxidative species have also been assessed indirectly by the presence of oxidative damage to DNA and
proteins as well as peroxidation of lipids. Studies assessed in the 2013 Pb ISA showed that Pb exposure
increased signs of oxidative damage in the brains of a variety of animal species (U.S. EPA, 2013; Wu et
al„ 2008). Since publication of the 2013 Pb ISA, more recent studies have demonstrated Pb-induced
increases in ROS production and oxidative damage in the brain both during development (Hossain et al„
2016; Lu et al„ 2013) and postweaning (Singh et al„ 2019; Liu et al., 2018a; Thangarajan et al„ 2018;
Singh et al., 2017; Kumar and Muralidhara, 2014; Flora et al., 2012). Proper regulation of oxidative stress
requires a balance between the presence of oxidative species (i.e., ROS and RNS) and levels of
antioxidant defense proteins (e.g., glutathione [GSH], catalase [CAT], and SOD). Along with increased
ROS/RNS production, depletion of antioxidant proteins or reductions in antioxidant enzyme activity
could contribute to an overall increase in oxidative stress. As discussed in the 2013 Pb ISA, animal
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studies and human panel studies have shown that BLL is associated with an increased ratio of oxidized to
unoxidized GSH (Mohammad et al., 2008; Diouf et al., 2006; Ercal et al., 1996; Sandhir and Gill, 1995).
More recent studies showed similar impairment of antioxidant defenses in animal models of Pb exposure
during developmental (Lu et al., 2013) and postweaning (Singh et al., 2019; Thangarajan et al., 2018;
Singh et al., 2017; Flora et al., 2012) Pb exposures. Furthermore, changes in antioxidant status and
oxidative stress can contribute to mitochondrial dysfunction, as described above. There is strong evidence
that Pb exposure across the lifespan disrupts multiple aspects of energy metabolism and oxidative stress
regulation.
Cellular damage caused by oxidative insults can trigger inflammation and vice versa; thus, it is
often difficult to disentangle which occurs first. Given the interrelated nature of these factors, they are
combined within the same gray box in the blood Pb diagrams (Figure 3-1 and Figure 3-2). Inflammation
is a hallmark of many neurological conditions and neurodegenerative diseases. Inflammation can be
triggered by the production of inflammatory mediators (e.g., cytokines) in response to cell or protein
damage. As discussed in the 2013 Pb ISA, Pb exposure results in signs of inflammation including
activation of inflammatory signaling pathways, inflammatory mediator production, and microglia cell
activation (U.S. EPA, 2013). Several studies have observed increased inflammatory mediator levels and
activation of inflammatory signaling pathways (for example, tumor necrosis factor-alpha) in the brains of
animals exposed to Pb during development (Chibowska et al„ 2020; Hossain et al„ 2016; Ashok et al„
2015) and postweaning (Yang et al„ 2019; Liu et al„ 2018a). Proinflammatory markers interact with, and
in some cases infiltrate, the BBB, initiating neuroinflammation, as indicated by altered gene expression,
increased apoptosis, lipid and protein oxidation, and microglial activation (Saleh et al., 2018; Shvachiy et
al„ 2018; Sobin et al., 2013). Similarly, histologic and immunohistochemical signs of neuroinflammation
in the dentate gyrus have been reported in rats exposed to Pb continuously from 7 days postconception to
28 weeks of age, which corresponded to behavioral changes (Shvachiy et al., 2018). In the same study, a
similar neuroinflammatory phenotype was observed in mice that were given an 8-week Pb abstinence
period between 12-week and 8-week Pb exposures (Shvachiy et al., 2018). In sum, recent evidence
supports the plausibility of inflammation as an intermediate event in the development of neurological
health effects regardless of the timing of Pb exposure.
While a robust immune response can protect the brain from certain insults, prolonged
neuroinflammation is associated with several neurological and neurodegenerative diseases. AD,
characterized by the accumulation of A|3 and p-tau, has been associated with increased markers of
neuroinflammation. While neurodegenerative diseases are associated with old age, studies of
developmental exposures to Pb have shown that early life exposures are associated with Alzheimer" s-like
pathology in adult animals. As discussed in the 2013 Pb ISA, Pb exposure in juvenile animals resulted in
the increased production of APP and higher levels of p-tau in offspring. Similarly, early life Pb exposure
of nonhuman primates led to Alzheimer"s-like pathology later in adulthood (Wu et al., 2008). Studies
published since the last ISA support and extend the findings that developmental exposures to Pb can lead
to increased levels of misfolded proteins (e.g., abnormal APP processing, A|3, tau protein) and
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Alzheimer"s-like pathologies (e.g., p-tau accumulation) (Ashok et al., 2015; Bihaqi and Zaw ia. 2013).
Evidence from exposures during development suggests that early life may represent a sensitive window
for insults associated with neurodegenerative disease (Liu et al.. 2014a). Some studies with exposure of
postweaning animals to Pb have shown increased inflammation associated with AD markers (Yang et al..
2019; Liu et al., 2018a; Zhang et al., 2012). In postweaning studies, treatment with molecules with anti-
inflammatory and antioxidative properties were able to prevent A|3 accumulation and reversed cognitive
and behavioral alterations in Pb-exposed mice (Liu et al., 2020; Yang et al., 2019; Liu et al., 2018a).
Because of this, there is a solid line connecting the box containing inflammation, oxidative stress, and
altered energy metabolism to the accumulation of A|3 in Figure 3-2. Evidence of effects of Pb on other
neurodegenerative diseases are more limited. A recent study showed that exposure of postweaning rats to
Pb resulted in increased accumulation of a-synuclein, a protein associated with PD, in the hippocampus
that correlated with impaired learning and memory (Zhang et al., 2012). Overall, new data support the
previous findings that Pb exposure can affect the development and progression of neurodegenerative
pathologies in developmentally and postweaning exposed animals.
Beyond Pb's ability to produce neuroinflammation and oxidative stress and increase expression
of disease-related proteins, excessive damage to cellular proteins or DNA can trigger cell death. While
cell death and neuronal population loss in adulthood contribute to brain pathology, the developing brain is
far more sensitive to disruption. Cell migration, differentiation, and pruning are all essential
neurodevelopmental processes that need to be carefully timed and orchestrated. Thus, increased or
aberrant cell loss results in improper nervous system development that could be responsible for the altered
mood, sensory, or cognitive functions observed in Pb-exposed children and animals. The 2013 Pb ISA
and Section 3.4.2.1 present several animal studies showing upregulation of apoptotic markers in various
regions of the brain following Pb treatment, at various lifestages, which was supported by similar findings
in in vitro experiments (U.S. EPA, 2013). Recent studies also reported activation of pro-apoptotic
pathways in response to developmental Pb exposure (Ebrahimzadeh-Bideskan et al., 2016; Hossain et al.,
2016; Su et al., 2016; Lu et al., 2013), supporting and extending the experiments reviewed in the 2013 Pb
ISA. Another study showed histologic changes in the brain concomitant with increased markers of protein
and lipid damage (Saleh et al„ 2019), suggesting a relationship between cell death and structural changes
in the brain with oxidative damage. Developmental Pb exposure caused dysregulated myelination in the
brains of rats, which could be rescued with cotreatment with antioxidants (Nam et al., 2020; Nam et al.,
2019a). Myelination is an essential step in nervous system development, as myelin sheaths facilitate quick
and efficient electrical transmission along nerve cells to preserve nervous system function and
connectivity. Several studies have also demonstrated that treatment with compounds with antioxidant
capacity reduced the apoptotic signaling (Nam et al., 2018b; Ebrahimzadeh-Bideskan et al., 2016).
Together, these data provide the justification for a solid line from the gray box containing oxidative stress
and inflammation to the box containing cell injury/death in Figure 3-2.
Widespread cell loss in the mature nervous system can also lead to functional and structural
changes that can contribute to behavioral and cognitive changes. Cell death is also a common element in
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many neurodegenerative diseases. The animal studies showing upregulation of apoptotic markers in
various regions of the brain following Pb treatment discussed in the 2013 Pb ISA are strengthened by
similar findings in several new studies (Amedu and Omotoso. 2020; Liu et al., 2020; Abubakar ct al..
2019; Singh et al., 2019; Yang et al., 2019; Liu et al., 2018a; Thangarajan et al., 2018; Maiti et al., 2017;
Singh et al., 2016; Flora et al„ 2012). In vitro exposure of neuronal cell lines to Pb resulted in reduced
cell viability and increased apoptosis (Ye et al., 2020; Liu et al., 2017; Neelima et al„ 2017; Meng et al„
2016; Su et al., 2016; Ye et al„ 2016b; Ahmed et al„ 2013). Additional discussion of apoptotic markers
and brain structural changes following Pb exposure are discussed in Section 3.4.2.1. Like the
developmental exposure studies, demyelination was observed in the spinal cord following postweaning
exposure to Pb (da Silva et al., 2020; Villa-Cedillo et al„ 2019). These data provide plausibility that adult
Pb exposures contribute to cognitive and behavioral changes. Some studies therapeutically targeted RNS
production in the mitochondria by treatment with fisetin, a polyphenolic compound with antioxidant
properties to ameliorate the activation of pro-apoptotic signaling (Yang et al„ 2019; Maiti et al., 2017).
Their results suggest a role for oxidative stress in triggering the apoptotic cascade. In vitro treatment with
the antioxidant genistein also protected against cell death (Su et al., 2016). Thangarajan et al. (2018)
showed that treatment with an anti-inflammatory and antioxidative compound, morin, was able to largely
restore proper brain architecture after Pb exposure. Together, these data provide the justification for a
solid line from the gray box containing oxidative stress and inflammation to the box containing cell
injury/death in Figure 3-2.
Inflammation and oxidative stress can also affect the integrity of the BBB, which provides a
selective barrier for entry from the circulation to the brain and spinal cord. As discussed in the 2013 Pb
ISA and 2006 Pb AQCD, Pb exposure in rodents was shown to increase permeability of the BBB and the
blood-CSF barrier. Interestingly, Pb alone does not have a large effect on BBB integrity but can prolong
BBB permeability in response to other stimuli (U.S. EPA, 2013, 2006). The effect of Pb on the BBB is
also selective in that the permeability of all solutes is not affected equally (U.S. EPA, 2013, 2006).
Disruption of the BBB could potentially promote increased Pb accumulation in the brain with prolonged
or repeated exposure. Two new studies assessed the integrity of the BBB following Pb exposure and
observed disruption of brain permeability with reduced levels of tight junction proteins and other
important capillary proteins (Wu et al., 2020a; Song et al., 2014). In adult rats, 8 weeks of Pb dosing
reduced expression of the tight junction proteins occludin and zonula occludens-1 at the BBB (Song et
al., 2014). Pb has been implicated in alteration of the CSF barrier in rats (Zheng et al., 1996). The CSF
can carry hormone signals important for brain development; thus, disruption of the cerebrospinal barrier
could affect proper hormone signaling for brain development. Indeed, Pb exposure was reported to
decrease transthyretin levels in the CSF, suggesting altered cerebrospinal barrier integrity (Zheng ct al..
1996). There is likely interplay between Pb effects on endocrine and nervous system development. In
conclusion, there is a potential for Pb to affect the BBB and blood spinal cord barrier, which could alter
Pb availability and uptake into the nervous system.
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While most data suggest that Pb acts through a mode of action involving oxidative stress and
inflammation, additional signaling pathways are also affected by Pb exposure. Pb exposure can alter ion
balance, which has particular importance with regard to the effect on Ca2+ signaling. Calcium signaling is
vital for many fundamental neurological processes including membrane excitability, neurotransmitter
release, synaptogenesis, transmission, and other processes. Of particular relevance for this review,
neurotransmitter signaling is intimately connected with Ca2+ signaling. As discussed in the 2013 Pb ISA,
developmental exposure to Pb interferes with the evoked release of neurotransmitters by inhibiting Ca2+
transport through voltage-gated ion channels (Cooper and Manalis, 1984; Suszkiw et al., 1984). Ca2+ is
also a ubiquitous second messenger, which can regulate many neuronal physiologic processes like gene
expression, membrane excitability, and dendrite development (Kawamoto et al., 2012). Interestingly, in
the absence of stimulation, Pb has some Ca2+ mimetic activity that increases baseline neurotransmitter
release (Cooper and Manalis, 1984; Suszkiw et al., 1984). In general, Pb exposure increased Ach levels,
increased dopaminergic signaling, and reduced NMDAR expression (U.S. EPA, 2013). Animal models
suggest that Pb exposure during development leads to inhibition of acetylcholinesterase (AchE), thereby
increasing the levels of Ach and causing lasting neurodevelopmental changes that persist into adulthood
(Basha and Reddy, 2015). The authors found that addition of Ca2+ restored cholinergic signaling (Basha
and Reddy, 2015). These data help to justify the solid line from ionic mimicry to altered neurochemical
signaling in Figure 3-1. Similar studies in animals postweaning have shown similar decreases in AchE
activity (Galal et al., 2019; Okesola et al., 2019; Thangarajan et al„ 2018; Andrade et al„ 2017; Ferlemi et
al„ 2014; Phyu and Tangpong, 2013). Recent literature also supports altered dopaminergic signaling
following postweaning Pb exposure (Sobolewski et al., 2020; Yousef et al„ 2019; Amos-Kroohs et al..
2016; Stansfield et al., 2015; Basha et al„ 2014; Weston et al., 2014; Cory-Slechta et al„ 2012; Graham et
al„ 2011). Ca2+ gradients are also responsible for generating action potentials. Alteration of intracellular
Ca2+ levels in neurons could cause deleterious effects on action potential generation and repolarization.
Recent evidence shows that hippocampal slices from 50-day old rats exposed to Pb both pre and
postnatally had enhanced pared pulse facilitation, suggesting Pb-induced dysregulation of Ca2+ signaling
(Zhang et al., 2015b). This data support findings from a limited number of human MRI and MRS studies
that provide evidence of physical and physiological changes in the brain corresponding to increased blood
Pb that were discussed in the 2013 Pb ISA and 2006 Pb AQCD. Together there is evidence that Pb can
alter neurotransmitter release and signal potentiation, which in turn could contribute to the changes in
brain activity seen in various behavioral and cognitive diseases.
Beyond the actions of Pb discussed thus far, there is growing evidence of the effect of changes to
the epigenome in mediating toxicity. Epigenetic changes refer to alterations in the mechanisms that
regulate gene expression without altering DNA sequence. Epigenetic programming is a fundamental
developmental process, and the complex relationships between the genome, epigenome, and environment
can shape the health of present and future generations. Epigenetic alterations are often measured as
changes in histone and DNA methylation patterns as well as the levels of the enzymes (e.g.,
methyltransferases, acetylases, deacetylases) responsible for regulating epigenetic modification in situ.
Transcriptional regulators like microRNAs and long noncoding RNAs are also considered epigenetic
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modifiers. As discussed in the 2013 Pb ISA, developmental Pb treatment in mice and monkeys decreased
the activity of some DNA methyltransferases. Therapeutic treatment with a methyl donor improved Pb-
induced decrements in LTP and Morris water maze performance (Cao et al.. 2008). Gestational and
postnatal exposure to Pb in rats increased histone acetylation in the hippocampus, which corresponded to
a hyperactivity phenotype (Luo et al.. 2014). The authors suggested this was due to upregulation of
histone acetyltransferases, including p300. However, these changes occurred at BLLs in excess of 50
(ig/dL; thus, the relevance of these findings to ambient exposure in humans is questionable. Other studies
have reported changes in the expression of DNA methyltransferases (Schneider et al.. 2013) and increased
hypermethylation, especially in the hippocampus of female mice (Sanchez-Martin et al.. 2015). The
window of exposure, prenatal stress, and sex can all play a role in determining the epigenetic
modifications (Sobolewski et al.. 2018). While differences in epigenetic modifications and the effects of
Pb on epigenetic enzymes have been reported and linked to behavioral effects in animals, there remains
little evidence to connect epigenetic changes to alterations in specific pathways that have the potential to
cause neurobehavior effects. As a result, the arrow for epigenetic changes in is represented as a dotted
line when connecting to the box for neurodevelopmental disorders. Future research may elucidate a role
for epigenetic modification in the etiology of neurological diseases. Most of the Pb literature on
epigenetic changes has focused on heritable epigenetic changes during development; however, the effect
of postweaning exposure to Pb on epigenetic mechanisms is not well known. Very few studies have
evaluated epigenetic changes throughout the lifetime or with later life exposures. Individual studies have
found changes in the expression of a long noncoding RNA (Nan et al.. 2016) and a methyltransferase
(Schneider et al.. 2012). which might suggest that epigenetic modification could be affected during adult
exposures; however, there are too few studies to draw reliable conclusions. The conclusion of epigenetic
modification resulting in neurological effects is only plausible for developmental exposures with the
present available data.
In summary, Pb exposure can result in a range of neurocognitive and behavioral health effects
through a myriad of complex biological pathways. The pathways described here provide biological
plausibility for associations between Pb exposure and nervous system effects in in children and adults.
The developmental timing, sex, and presence of other stressors or enrichments alongside of Pb exposure
can affect the resulting health effects. The identified pathways share many common features, including Pb
interactions with cellular proteins, competing with and displacing other biologically relevant cations,
increased oxidative stress, and inflammation, which can have widespread effects on brain structure and
function. There is also evidence for disruptions of Ca2+ signaling, which can result in altered
neurotransmitter signaling and contribute to the development of neurological health effects. Epigenetic
modifications resulting from Pb developmental exposure have been reported but are still an area of active
investigation. The role of these epigenetic changes in the progression of neurological health effects with
later Pb exposures is unclear. Together the proposed pathways provide biological plausibility for
epidemiologic evidence of neurological effects and were used to inform causality determinations
throughout this appendix.
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3.4
Overt Nervous System Toxicity
Overt nervous system toxicity refers to a diverse group of endpoints that inform brain structure
and function, including brain histopathology changes, brain weight, electrophysiology,
neuroinflammation, and neurotransmitter analyses. The collective body of epidemiologic and
experimental animal studies assessed in the 2013 Pb ISA demonstrated the effects of Pb exposure on an
array of nervous system outcomes. The evidence, including uncertainties, is summarized in Section 3.4.3.
Study details that supplement the information provided in the text are in the evidence inventories (Table
3-IE and Table 3-IT in Section 3.7). Previous Pb assessments reviewed epidemiologic studies that found
associations of Pb biomarkers with electrophysiologic or physical changes in the brains of adults assessed
by imaging technologies. Biological plausibility for the effects of Pb on overt nervous system toxicity
was provided by a small number of experimental animal study findings with dietary and lactational Pb
exposure, with some evidence at BLLs relevant to humans. Recent epidemiologic studies support and
extend the evidence pertaining to the association of lead exposure during childhood with brain structure
and function in adolescence or adulthood. A smaller set of cross-sectional studies also report associations
between childhood BLLs and overt nervous system outcomes. Multiple experimental animal studies
report changes in the brains of rats and mice following exposure to Pb. These include changes in
histology, neurotransmitter measures, brain weight, and electrophysiology measures, providing coherence
for the epidemiologic studies that show associations with decrements in cognition, neurodegeneration, or
increased behavioral problems. There is no causality determination for this section; rather, evidence in
this section may be referenced in the outcome-specific "Summary and Causality Determination"
discussions of Sections 3.5 and 3.6 if they provide biological plausibility or coherence for the
observations in the epidemiologic studies.
3.4.1 Epidemiologic Studies of Brain Structure and Function
Previous Pb assessments (U.S. EPA, 2013, 2006) reviewed a small body of epidemiologic
studies that found associations of Pb biomarkers with electrophysiologic and physical changes in the
brains of young adults as assessed by magnetic resonance imaging (MRI) or spectroscopy (MRS). The
implications of findings from most studies assessed in the 2006 Pb AQCD were limited by the small
sample sizes (n = 12 to 45) and inadequate consideration of potential confounding. However, analyses
cohort of adults (ages 20-23 years) reviewed in the 2013 Pb ISA included larger sample sizes and aimed
to characterize potentially important lifestages of Pb exposures (Yuan et al„ 2006), thus expanding the
evidence pertaining to potential links between physiologic brain changes and functional
neurodevelopmental effects. Overall, the small number of studies in a limited number of populations
assessed in the 2013 Pb ISA showed physical and physiologic changes in areas of the brain associated
with neurodevelopmental function, providing biological plausibility for the associations
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observed between Pb biomarker levels and cognitive function decrements and behavioral problems.
Several recent longitudinal studies add to the evidence characterizing the association of Pb
exposure during childhood with brain structure and function during adolescence or adulthood. A smaller
number of studies evaluated the cross-sectional association of childhood BLL with brain structure or
function in childhood. These studies are summarized below, and key information from the studies is
included in Section 3.7, Table 3-IE.
Reuben et al. (2020) conducted a study to examine the effect of childhood BLL, measured at age
11, on lower structural integrity of the brain at age 45. These investigators used data from the Dunedin
Study in New Zealand, which enrolled participants beginning in 1972 and 1973 and followed them
through April 2019. The mean early childhood BLL for the study participants was 10.99 (ig/dL. MRI was
used to assess multiple endpoints related to gray matter (cortical thickness, surface area, and hippocampal
volume), white matter (white matter hyperintensities, fractional anisotropy [FA]), and the gap between
chronological age and estimated brain age. In addition, cognitive function was estimated using the
Wechsler Adult Intelligence Scale (WAIS)-IV, self-reports, and informant reports (see Section 3.6.1). A
total of 564 of the original 1037 infants enrolled at birth were included in the analysis. Findings from the
study are depicted in Figure 3-3. In models adjusted for sex, maternal IQ, and socioeconomic status
(SES), associations were observed with cortical surface area, hippocampal volume, global FA, and the
gap between each study member's chronological age at imaging and their MRI-predicted age, but not
with all the MRI metrics assessed.
Two analyses of the Cincinnati Lead Study (CLS) have been conducted since the 2013 Pb ISA.
Confounders considered in these analyses included child characteristics, Home Observation for the
Measurement of Environment (HOME) score, maternal IQ, and SES (see Section 3.7, Table 3-IE for
study-specific confounders). Cecil (2011) examined the association of childhood BLL (childhood [3-28
months] average) with volumetric MRI, MRS, diffusion tensor imaging (DTI), and functional MRI
outcomes ascertained between ages 19 and 24 years old. This study found that childhood BLL was
associated with decreased gray matter volume in several regions (i.e., medial and superior frontal gyri,
inferior parietal lobule and cerebellar hemispheres). Higher childhood BLL was also associated with
lower metabolite concentrations in several brain regions (white matter, left basal ganglia, left cerebellar
hemisphere, and vermis). DTI and functional MRI findings also suggested injury and compensatory
activity in specific brain regions. Overall, structural, organizational, and functional changes in the brain
regions responsible for regulating behavior were indicated by this study. In another study of participants
enrolled in the CLS, Beckwith et al. (2021) examined the relationships between childhood BLL (at 78
months), structural brain volume, and adult criminality. BLLs were associated with MRI-derived
decreases in white and gray matter volumes in the frontal parietal and temporal lobes. Decreased gray
matter volume in brain regions responsible for cognition and emotional regulation was also associated
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with criminal arrests, potentially supporting associations observed between Pb exposure and conduct
disorders that are described in Section 3.5.3.
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WMH = white matter hyperintensities; yr = year(s).
Regression lines and their 95% CIs are plotted. The box plots show the distribution of BLLs and brain outcomes.
Beta coefficients shown are for an incremental increase of 5 (jg/dL in childhood BLL. Source: Reuben et ai. (2020).
Figure 3-3 The relationship between blood Pb level at age 11 and brain
outcomes in adulthood.
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Lamoureux-Tremblav et al. (2021) examined the association between pre- and postnatal
(measured concurrently to the MRI) BLLs and MRI findings in adolescents (mean age 18.3 years)
enrolled in a longitudinal study of Inuit from Northern Quebec exposed to Pb, mercury (Hg), and
polychlorinated biphenyls (PCBs). Functional MRI data were collected during fear conditioning and
extinction tasks with the aim of understanding emotional dysregulation that could lead to anxiety
disorders. These authors found higher differential activation in the right dorsolateral prefrontal cortex in
association with higher postnatal BLL. Differential effects in the high Pb exposure group were observed
during the fear extinction phase and maintained discrimination between the safety and threatening signals
(CS+ > CS-). Activation of the dorsolateral prefrontal cortex has been associated with cognitive processes
for regulating the affective state. The mean cord blood Pb was 4.56 (ig/dL and the mean concurrent BLL
was 1.78 (ig/dL in this study. In an earlier study of this Inuit population, Ethier et al. (2012) measured
visual evoked potentials (VEPs) using electrodes on the scalp to provide a direct measure of brain
functions related to sensory function (i.e., visual contrast sensitivity and spatial vision) at age 5. The mean
cord BLL in this study was 4.6 (ig/dL. Amplitude and latency for standard VEP components were
measured (i.e., N75 [negative deflection at approximately 75 ms], P100 [positive deflection at
approximately 100 ms], and N150 [negative deflection at approximately 150 ms]) in this longitudinal
study. Multiple tests were conducted, and associations were reported with significance levels. Cord Pb
level was associated with a delay of the N150 component (e.g., |3 = 0.06 [95% CI: 0.01, 0.10]) and other
latency metrics, which may indicate a deficit in early visual processing in Pb-exposed children.
Confounders including child characteristics, maternal education, SES, drug and alcohol use, and other
metals were considered in the analyses of Inuit children (see Table 3-IE for study-specific confounders
considered). A subset of two separate studies was combined for this study, and participation rates based
on the original number of participants enrolled in the study were not reported.
In a cross-sectional analysis of children, Kim et al. (2018a) examined the interaction between
dopamine receptor D2 (DRD2) and BLL on the cortical thickness of 12 regions of the frontal lobe
ascertained via MRI. The D2 receptor is located in the prefrontal cortex of the brain and may contribute to
the pathology of attention deficit/hyperactivity disorder (ADHD). The authors relied on an age- and sex-
matched sample of children ages 6 to 17 years old with and without confirmed ADHD for the analysis.
This study found an interaction effect between a variant of DRD2 and BLL on reduced cortical thickness
of several regions in the frontal lobe in the ADHD group, but not in the healthy controls in regression
analyses adjusted for age, intracranial volume, and sex. A correlation between reduced cortical thickness
and poorer inattention score on the parent-reported ADHD rating scale was also reported, supporting a
link between Pb exposure, MRI findings, and functional decrements in attention. Results for Kim et al.
(2018a) are found in Section 3.7, Table 3-IE.
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3.4.1.1 Summary
Overall, multiple studies, including prospective studies following children through adolescence
and adulthood, found associations between BLL and physiologic changes in regions of the brain
responsible for cognition and behavior. The prospective studies considered important confounders
including the HOME score (CLS only), SES, drug and alcohol use, and maternal IQ. A smaller number of
studies indicated Pb-related changes in brain function or physiologic changes in children. Associations
between Pb exposure and a large number of metrics and brain regions were evaluated in the
epidemiologic studies, raising the likelihood of chance findings.
3.4.2 Experimental Animal Studies of Brain Structure and Function
Animal studies can offer insights into Pb-induced effects on brain structure and function through
investigations that cannot be conducted on human subjects. Experimental animal studies provide evidence
that Pb can cause alterations in brain development. The 2013 Pb ISA reviewed evidence that Pb exposure
produced increased levels of oxidative stress and inflammatory response markers (U.S. EPA, 2013).
These effects were observed in many regions of the brain and were associated with changes in neuronal
and glial cell morphology, neurotransmitter levels, and brain electrophysiology. Additional discussion of
the biological pathways that potentially underlie these nervous system effects are discussed in Section 3.3.
Recent studies (see Figure 3-3 and Table 3- IT) support the results summarized in the 2013 Pb
ISA, showing increases in inflammatory responses and markers of oxidative stress in various regions of
the brain. Most studies evaluated oral dosing of Pb via drinking water or gavage with exposure durations
ranging from 14 days to 701 days depending on the study, with studies in both adult and developing
animals. The specific dosing regimen varied by study, but the present review focuses on studies that
resulted in a measured BLL <30 (ig/dL. In studies with multiple time points, the magnitude or severity of
effects generally increased with exposure duration. In addition, studies with developmental Pb exposure
identified pregnancy and early development as sensitive windows for Pb toxicity. Several studies have
examined brain architecture using histological methods following Pb exposure. In general, the magnitude
and severity of the effects increased with longer exposure durations, and some studies found Pb-induced
effects could be ameliorated by co-exposures with antioxidants (such as vitamins and specific lipids) or
by environmental conditions (e.g., rearing condition or enrichment).
3.4.2.1 Histopathology
The nervous system is made up of the central and peripheral nervous systems, which include a
diversity of cell types broadly grouped into neurons and glial cells. Neurons are the functional electrically
excitable cells in the brain. Glial cells can be further divided into microglia and macroglia (astrocytes,
oligodendrocytes, ependymal cells, Schwann cells, satellite cells, radial glia, and enteric glia), which are
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neuronal support cells with diverse functions including innate immunity, phagocytosis, myelination,
synaptic regulation, neuronal activity, and blood-brain barrier (BBB) integrity (Rea. 2015). In the central
nervous system (CNS), astrocytes, oligodendrocytes, and microglia are the major types of glial cells.
Normal brain development relies on coordination of all cell types across time. Pathology methods can be
used to study alterations in brain cell morphology, function, and composition.
Short-term studies (<30 days) in adults evaluating Pb exposure and brain morphology after oral
exposures in rats have observed abnormalities in treated animals and their offspring, including decreased
numbers of neurons or synapses (Nam et al.. 2018a; Saleh et al.. 2018; Gassowska et al.. 2016a; Han et
al.. 2014; Rahman et al.. 2012b). disorganized cells and lack of characteristic layering (Saleh et al.. 2019;
Saleh et al.. 2018; Zhou et al.. 2018). increased vacuolization (Saleh et al.. 2019). effects on dendritic
spines (Xiao et al.. 2020; Saleh et al.. 2018; Wang et al.. 2016; Du et al.. 2015; Rahman et al.. 2012b).
increased numbers of apoptotic cells (Saleh et al.. 2019; Meng et al.. 2016). and increased expression of
various proteins and biochemical parameters related to oxidative stress (Saleh et al.. 2018; Singh et al..
2017; Zhu et al.. 2013) in multiple brain regions including the cerebellum, cerebral cortex, and
hippocampus. In aggregate, this evidence suggests that Pb has the potential to disrupt the integrity of
single neurons and populations, which may contribute to overt toxicity. Pb exposure has also been shown
to interfere with the homeostasis of other essential metal ions, such as iron (Fe), in the brain (Zhu et al..
2013) and to inhibit various enzymes involved in energy production or glucose uptake (Zhao et al.. 2021)
and metabolism (reviewed in (ATSDR. 2020)). For all these studies, BLLs were below 30 (ig/dL, and
many were below 20 (ig/dL (see Evidence Inventory Table 3- IT). Further discussion of these mechanisms
is provided in Section 3.3.
Studies of long-term Pb exposure (>30 days) have also assessed brain morphology in adult mice
and rats following oral dosing and observed neuronal damage including irregular shape, vacuolization,
and cell degeneration in the cerebellum, hippocampus, and cerebral cortex (Liu et al.. 2022c; Saleh et al..
2019; Singh et al.. 2019; Saleh et al.. 2018; Sun et al.. 2014). Other histopathological lesions, including
karyopyknosis (pre-apoptotic chromatin condensation of cell nuclei) (Nan et al.. 2016) and swollen and
distorted mitochondria (Ouvang et al.. 2019; Gassowska et al.. 2016a; Sun et al.. 2014) were observed in
other studies. Several biochemical parameters related to oxidative stress were assessed including
apoptosis using immunohistochemical methods like terminal deoxynucleotidyl transferase dUTP nick end
labeling (TUNEL staining) (Singh et al.. 2019; Baranowska-Bosiacka et al.. 2017; Meng et al.. 2016; Nan
et al.. 2016; Su et al.. 2016; Baranowska-Bosiacka et al.. 2013) as well as caspase-3 (Wang et al.. 2021a)
and cell replication using proliferating cell nuclear antigen (PCNA) (Singh et al.. 2019). In a 12-week rat
dietary study, the authors reported neuronal damage and cognitive deficits accompanied by decreased
levels of synaptic proteins as well as decreased levels of receptors and proteins related to synaptic
plasticity regulation (N-methyl D-aspartate receptor [NMDAR], cyclic adenosine 3",4"-monophosphate
response element binding protein [CREB], brain-derived neurotrophic factor [BDNF]) (Liu et al.. 2022c).
BLLs in all of these studies were below 30 (ig/dL and many were below 20 (ig/dL (see Evidence
Inventory, Table 3-IT).
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In the single available inhalation study, adult mice were dosed for 6 weeks via whole body
inhalation of Pb oxide (PbO) nanoparticles. Hippocampal damage was observed including shrunken and
damaged neurons, as well as increased Pb content in the brain (Dumkova et al.. 2017). The study reported
BLLs of 13.99 (ig/dL. However, exposure to PbO nanoparticles via inhalation did not impact markers of
cell proliferation (PCNA) or cellular apoptosis (TUNEL) in the hippocampus (Dumkova et al.. 2017).
Despite the limited evidence available for inhaled Pb, continuity of the effects (i.e., neuronal damage) has
been reported across different routes of exposure. Toxicology studies of Pb inhalation remain a data gap
that limits the evaluation of nervous system effects from inhaled Pb.
Brain morphology has also been assessed in developing animals with a variety of exposure
contexts including pre-mating, during gestation, and during lactation. Pb is transferred across the placenta
and through lactation (Silbergeld. 1991; Bhattacharvva. 1983). Some studies included cross fostering
experiments to assess unique sensitivities at specific developmental windows, with some evidence
indicating that the postnatal period is particularly sensitive to Pb neurotoxicity (Barkur and Bairv. 2016).
Given the altricial nature of rodents, the postnatal period is roughly analogous to the third trimester of
human brain development. Similar to results in adults, studies of developmental Pb exposures observed
alterations in brain morphology including reduced numbers of neurons in the forebrain, hippocampus,
hypothalamus, and amygdala (Long et al.. 2022; Vigucras-Villascnor et al.. 2021; Wang et al.. 2021a;
Nam et al.. 2018a; Shvachiv et al.. 2018; Xiao et al.. 2014). damaged neurons (Long et al.. 2022; Wang et
al.. 2021a; Zhu et al.. 2013). reduced numbers of glial cells (Dominguez et al.. 2019; Sobin et al.. 2013).
changes in synapses (Sadeghi et al.. 2021; Wang et al.. 2021b; Gassowska et al.. 2016a; Gassowska et al..
2016b; Zhang et al.. 2015b; Xiao et al.. 2014). reduced numbers of mitochondria (Zhang et al.. 2015b).
swollen and shrunken mitochondria (Ouvang et al.. 2019; Gassowska et al.. 2016a; Baranowska-Bosiacka
et al.. 2013). altered levels of glycoconjugates (constituents of synaptic and neural membranes) (Sadeghi
et al.. 2021). and chromatin abnormalities (Ouvang et al.. 2019; Baranowska-Bosiacka et al.. 2013). BLLs
in all of these studies were below 30 (ig/dL and many were below 20 (ig/dL. In a rat developmental study
in which animals were dosed throughout pregnancy and lactation (gestational day [GD] 1 to postnatal day
[PND] 21) with BLLs of 6.86 (ig/dL, the authors reported pathological changes in synapses, including
swelling of nerve endings, thickened synaptic cleft structure, and abnormalities in synaptic vesicle density
(Gassowska et al.. 2016a). These changes in synapse morphology were accompanied by decreases in key
synaptic proteins, as well as BDNF, a key neurotrophic factor that supports the differentiation,
maturation, and survival of neurons in both development and adulthood (Gassowska et al.. 2016a).
Additionally, studies in adult humans suggested that decreases in BDNF are associated with
neurodegenerative diseases (Bathina and Das. 2015). These changes in synapses can result in synapse
dysfunction, which would contribute to altered neurotransmission. Adverse changes in neuronal dendrite
morphology were reported following developmental Pb exposures in multiple studies, including loss of
dendritic spines, reduced spine density, decreased spine length, and impaired spine maturity and
morphology at multiple developmental stages and brain regions (hippocampus, medial prefrontal cortex,
dentate gyrus) (Xiao et al.. 2020; Saleh et al.. 2018; Zhao et al.. 2018; Sepehri and Ganji. 2016; Wang et
al.. 2016; Du et al.. 2015; Rahman et al.. 2012b). Dendritic spines are the morphological and structural
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basis for synaptic plasticity, learning, and memory (Frank et al.. 2018). providing biological plausibility
for altered learning and memory, as reported in Section 3.5.1.3.2.
The brain and CNS are separated from the blood by the BBB and the blood-cerebrospinal fluid
(CSF) barrier, which allows for selective transport of materials into the CNS. The BBB is formed by
several cell types including endothelial cells, astrocytes, pericytes, and microglia and plays an important
immunological role in protecting the brain from circulating pathogens and toxic substances. Two studies
assessed the integrity of the BBB following Pb exposure and observed disruption of brain permeability
with reduced levels of tight junction proteins and other important capillary proteins (Wu et al.. 2020a;
Song et al.. 2014). Increased permeability of the BBB may exacerbate neurotoxicity, as more toxicants
can penetrate the brain with repeated or continuous exposure. Additional research is needed to fully
elucidate the effects of Pb exposures on the BBB and the potential implications for CNS function and
disease.
Neuroinflammation is a complex response that involves microglia and astrocyte activation, as
well as other signaling proteins and cells (such as cytokines, reactive oxygen species, decreased
antioxidant activity). Inflammation is protective against pathogens but can result in neuronal injury or
neuronal loss in the CNS. Several studies in rodents evaluated and observed an effect of Pb exposure on
markers of neuroinflammation, including microglia and astrocyte activation, and the promotion of cellular
reactivity and inflammation (Wu et al.. 2020a; Saleh et al.. 2018; Shvachiv et al.. 2018). Numerous
studies also reported decreased numbers of microglia, which are glial cells that function primarily as
immune cells with macrophage activity, clearing cellular debris and dead neurons from nervous system
tissue (Dominguez et al.. 2019; Sobin et al.. 2013). Reduced numbers of microglia could indicate reduced
capacity for clearing cellular debris and responding to pathogens, which could contribute to functional
and morphological brain changes.
Experimental animal studies of rodents have also shown that Pb exposures affect measures of
brain metabolism, including reduced glycogen concentrations in various brain regions (such as the
forebrain, hippocampus, and cerebellum) (Baranowska-Bosiacka et al.. 2017) and reduced rates of
metabolism, which could indicate reduced glucose availability and poor metabolic cooperation between
neurons and astrocytes (Baranowska-Bosiacka et al.. 2017). A recent study directly measured decreased
hippocampal glucose metabolism following developmental Pb exposure (pre-mating through PND 10)
through reductions in glucose transporters (Zhao et al.. 2021). Glucose cannot be synthesized or stored in
neurons, hence glucose supply and transport are essential for neurophysiological processes with high
glucose demands (such as learning and memory). Because the brain has the highest energy demand and
metabolism of any organ, energy homeostasis is critical. In the above study, effects on glucose
metabolism persisted at PND 30 when the blood Pb concentration had returned to control levels (BLL
was 11.4 (ig/dL at PND 10 and 1.8 (ig/dL BLL at PND 30).
Taken together, animal studies provide strong evidence that Pb exposure impacts brain structure
and function. Altered brain morphology, increased brain inflammation, oxidative stress, and associated
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mitochondrial damage have all been consistently reported following Pb exposures. These effects were
observed across multiple brain regions, on different levels of brain organization, across a variety of
lifestages, and in both sexes.
3.4.2.2 Neurotransmitter Analysis
Neurotransmitters are molecules involved in the transmission of chemical signals between
neurons and target cells and are involved in controlling a wide variety of brain functions, including motor
function, learning, memory, metabolism, behavior, and hormone production. These neurochemical
systems have been implicated in the initiation and maintenance of some brain diseases and disorders, e.g.,
Parkinson's disease (PD), depression, aggression, and dementia (Monday et al.. 2018; Chichinadze et al..
2011; Haden and Scarpa. 2007; Webster. 2001). As described in the 2013 Pb ISA, exposures to Pb can
induce changes in brain neurochemistry and signaling that vary by brain region, neurotransmitter type,
and the sex of the animal. Pb can compete with calcium ions (Ca2+) for common binding sites and second
messenger system activation. When Pb activates a Ca2+-dependent system in the nervous system, it can
contribute to spurious neurotransmitter regulation and release because this system intimately relies on
Ca2+ signaling for its homeostasis. Pb-related alterations in neurotransmission are discussed in further
detail below.
A variety of neurotransmitters and their metabolites were evaluated in experimental animal
studies of rodents, across multiple brain regions (hypothalamus, cerebral cortex, nucleus accumbens,
frontal cortex, striatum, hippocampus, olfactory bulb, midbrain, cerebellum) and time points, including
serotonin and its metabolite 5-hydroxylindolacetic acid (Weston et al.. 2014; Mansouri et al.. 2013;
Graham et al.. 2011). norepinephrine and its metabolite methoxyhydroxyphenylglycol (Long et al.. 2022;
Basha et al.. 2014; Weston et al.. 2014; Biioor et al.. 2012; Graham et al.. 2011). dopamine and its
metabolites dihydroxyphenylacetic acid and homovanillic acid (Sobolewski et al.. 2020; Amos-Kroohs et
al.. 2016; Stansfield et al.. 2015; Basha et al.. 2014; Weston et al.. 2014; Corv-Slechta et al.. 2012;
Graham et al.. 2011). acetylcholine (Long et al.. 2022; Mansouri et al.. 2013). epinephrine (Basha et al..
2014). glutamate and its precursor glutamine (Long et al.. 2022). The direction of changes depended on
the brain tissue analyzed, time point, sex, and specific neurotransmitter assessed. However, multiple
studies found significant effects of Pb exposure on the dopamine system (Sobolewski et al.. 2020; Amos-
Kroohs et al.. 2016; Stansfield et al.. 2015; Basha et al.. 2014; Weston et al.. 2014; Corv-Slechta et al..
2012; Graham et al.. 2011). Some studies also reported Pb-induced changes in enzymes involved in
neurotransmitter turnover and cycling, including monoamine oxidase (Basha et al.. 2014) tyrosine
hydroxylase (Sobolewski et al.. 2020). glutamine synthase, and adenylate cyclase (Long et al.. 2022).
Pb exposure has demonstrated effects on several neurotransmitters, which are important signaling
molecules that control multiple brain functions. These effects were observed across multiple brain
regions, across a variety of lifestages, and in both sexes. Altered neurotransmitter signaling can contribute
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to multiple brain dysfunctions and disorders, providing biological plausibility for the health effects
discussed in subsequent sections.
3.4.2.3 Brain Weight
Organ weights are a frequently assessed in experimental animal studies as they can easily be
measured during animal necropsy. Importantly, brain weight is an indication of severe toxicity as the
body goes to great lengths to spare the brain at the expense of other bodily systems. Brain weights were
not reviewed in the 2013 Pb ISA. Several recent studies of rodents assessed brain weight following short
and long-term Pb exposures, and most of these studies reported nonsignificant findings (Vigucras-
Villasenor et al.. 2021; Mani et al.. 2020; Wu et al.. 2020a; Singh et al.. 2019; Rahman et al.. 2018; Saleh
et al.. 2018; Zhou et al.. 2018; Singh et al.. 2017; Barkur and Bairv. 2015a; Wang et al.. 2013; Rahman et
al.. 2012b). However, two studies did find significant decreases in brain weight (16%—21% decrease): one
reported a decrease after a 90-day oral exposure to Pb in juvenile rats, which resulted in a BLL of 28.4
(ig/dL (Singh et al.. 2019) and the other reported an 18% decrease in cerebellum weight in treated dams
(27.7 (ig/dL BLL), as well as reduced fetal brain weight at parturition following gestational exposure to
Pb in drinking water (GD 1 to GD 20) (Saleh et al.. 2018). In addition to gross measures of brain size and
morphology (e.g., wet weight), studies using newer anatomical imaging methods have been conducted
since the 2013 Pb ISA. Three-dimensional imaging technologies, such as MRI and ultrasound methods
are being used for the analysis of neuroactivity and phenotypes in rodent toxicology studies (Turnbull and
Mori. 2007). Brain volume and MRI morphometry were assessed in a single study of developing mice
following dietary exposure to Pb (Abazvan et al.. 2014). The authors reported no changes in lateral
ventricle volume but did observe sex-specific changes in morphology including enlarged lateral
ventricles. As the technologies improve, imaging technologies offer promising results for evaluating
physiological functions like neural activity in whole intact animals. Study details are provided in Table
3-1T.
3.4.2.4 Electrophysiology
The effects of Pb exposure on brain electrophysiology were not reviewed in the 2013 Pb ISA.
Several new studies in rodents have found that Pb exposure affects measures of brain electrophysiology,
including long-term potentiation (LTP) and evoked excitatory postsynaptic currents (EPSCs). LTP is the
process of signal transmission by which synaptic connections between neurons are activated and
strengthened and may be one of the mechanisms underlying learning and memory processes. Recording
of LTP is a recognized model for the study of memory (Lynch etal.. 1990).
Presynaptic plasticity can be assessed using paired-pulse stimulation, wherein two stimuli occur
in close succession. Hippocampal slices from rats were subjected to LTP induction and high-frequency
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tetanic stimulations, and the magnitudes of EPSCs were measured. In developmentally Pb-exposed rats,
the magnitudes of EPSCs were lower at PND 10, suggesting impaired hippocampal induction (Zhao ct al..
2018). Other studies have also assessed EPSCs with a longer exposure duration and different conditions
and found the ratios of EPSC responses between paired-pulse stimuli were significantly greater in
hippocampal slices from Pb-exposed rats (Zhang et al.. 2015b). These changes in EPSC responses were
accompanied by inhibition of synaptic vesicular release (Zhang et al.. 2015b). Depressed LTP following
Pb exposures was measured in (Zhou et al.. 2020a; Wang et al.. 2016; Liu et al.. 2012). These studies
additionally reported increased neuronal free Ca2+ concentration and inhibition of various signaling
proteins (Ca2+ calmodulin dependent protein kinase II and CREB), which were mediated by upregulation
of the ryanodine receptor. Ryanodine receptors are ion channels that are critical for maintaining
intracellular Ca2+ homeostasis. Changes in electrophysiological parameters affect neurotransmitter release
and cell signaling, which can affect brain function (described in Section 3.4.2.2).
In addition to these effects, (Zhu et al.. 2019a) reported alterations in cardiac sympathetic nerve
activity in rats while evaluating nerve discharge as a potential contributor to other health effects discussed
in the cardiovascular toxicity section (Appendix 4). The authors reported enhanced cervical sympathetic
nerve discharge 1 year after Pb exposure ended, suggesting that Pb-induced alterations to autonomic
nervous dysfunction can have lasting effects. This growing area of research recognizes the potential effect
of Pb on electrophysiology in the nervous system.
3.4.2.5 Circadian Rhythms
Two recent studies evaluated the effects of Pb exposure on circadian rhythms using rodent
models. The suprachiasmatic nucleus (SCN) of the hypothalamus is the primary regulator of circadian
physiological processes and is synchronized daily by signals of light. Vigueras-Villasenor et al. (2021)
subjected male rats to chronic Pb exposure from conception to euthanasia. In these adult rats, under a
standard 12:12-hour light-dark cycle, the authors observed daily delays in the nocturnal onset of
locomotor activity. With a 6-hour photoperiod delay, the activity rhythms of Pb-exposed rats entrained to
a new cycle faster than controls, and Pb treatment showed no significant effects when the photoperiod
was advanced by 6 hours. Histochemical analyses of the hypothalamus in light-pulsed Pb-treated animals
displayed decreases compared with controls in both photo-stimulated neurons (immunoreactivity to c-
Fos) and the neuronal population in the SCN. Hsu et al. (2021) assessed disturbances in rodent sleep
homeostasis by using electroencephalography and electromyography to score the sleep wake architecture
of sleep cycles and found that adult rats with chronic Pb exposure showed disturbances in sleep patterns
that were accompanied by altered clock gene expression and changes in the hypothalamus. These
alterations in behavior, sleep cycles, brain structure, neuronal function, and gene expression warrant
further investigations into the effects of Pb on the rhythm of vital circadian processes. The above studies
reinforce the importance of considering the time of day in studies measuring the effects of Pb.
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3.4.2.6
Summary
In conclusion, multiple studies measured a variety of nervous system endpoints in brains of rats
and mice following exposure to Pb including histology, neurotransmitter analysis, brain weight, and
electrophysiology measures. Histological analyses revealed reduced neuron counts, altered synapse
morphology, and increased apoptosis, as well as oxidative damage in several brain regions, including the
hippocampus, frontal cortex, and cerebellum. These regions were also found to have damaged
mitochondria, vacuolization, and morphological changes. Pb concentrations ranged from 4.7 (ig/dL to
28.4 (ig/dL in these studies, which were conducted in a variety of animal models, sexes, and lifestages.
These endpoints provide biological plausibility for effects on cognitive behavioral changes and diseases
described in subsequent sections.
3.4.3 Integrated Summary of Overt Nervous System Toxicity
Overt nervous system toxicity refers to a diverse group of endpoints that inform brain structure
and function, including brain histopathology changes, brain weight, electrophysiology,
neuroinflammation, and neurotransmitter analyses. As described in Section 3.1, there are no causality
determinations for this endpoint grouping. Instead, the evidence is considered supporting information that
informs the health determinations in Sections 3.5 and 3.6.
Multiple studies measured nervous system endpoints in rats and mice following exposure to Pb
including histological changes in brain structure and morphology, neuroinflammation, neurotransmitter
analysis, brain weight, and electrophysiology measures. These findings that Pb exposures affect these
endpoints provide biological plausibility for Pb to elicit human cognitive behavioral changes and diseases
described in subsequent sections and are generally coherent with the epidemiologic studies of overt
nervous system effects described in this section. The lack of toxicology studies examining Pb inhalation
remains a data gap that limits the evaluation of nervous system effects from inhaled Pb.
Multiple epidemiologic studies, including prospective studies following children through
adolescence and adulthood and a smaller number of cross-sectional studies of children, found associations
between BLL and physiologic changes in regions of the brain responsible for cognition and behavior.
These prospective epidemiologic studies considered important confounders; however, many Pb exposure
metrics and brain regions were evaluated in the epidemiologic studies, raising the possibility of chance
findings.
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3.5 Nervous System Effects Ascertained during Childhood,
Adolescent, and Young Adult Lifestages
The collective body of epidemiologic and experimental animal studies assessed in the 2013 Pb
ISA demonstrated the effects of Pb exposure on an array of nervous system outcomes. Overall, the largest
body of evidence assessed in the 2013 Pb ISA, as well as the outcome that was best substantiated to occur
at the lowest Pb exposure levels, was related to cognitive effects in children. Multiple prospective studies
conducted in diverse populations consistently demonstrated associations of higher blood and tooth Pb
levels with lower full-scale IQ (FSIQ), executive function, and academic performance and achievement.
The blood Pb biomarkers used in these studies reflect exposures during prenatal, postnatal and childhood
lifestages. Tooth Pb generally reflects prenatal and early childhood exposure (or exposure up to the time
that the tooth is shed depending on the specific tooth layers analyzed) (see Section 2.3.4.1.) Most studies
examined representative populations and had moderate to high follow-up participation with no indication
of selective participation among children with higher BLLs and lower cognitive function. Associations
between BLL and cognitive function decrements were found with adjustment for several potential
confounding factors, most commonly SES, parental IQ, parental education, and parental caregiving
quality. In children aged 4-11 years, associations were found with prenatal, early childhood, childhood
average, and concurrent BLLs in populations with mean or group BLLs in the range of 2-8 (ig/dL.
Although examined less extensively than cognitive effects, a strong body of evidence also indicated Pb-
associated decrements in attention and increased hyperactivity.
3.5.1 Cognitive Function in Children
The evidence evaluated in the 2013 Pb ISA was sufficient to conclude that there is a "causal
relationship" between Pb exposure and decrements in cognitive function in children (U.S. EPA, 2013).
Multiple prospective studies conducted in diverse populations consistently demonstrated associations of
higher blood and tooth Pb levels with lower FSIQ, executive function, and academic performance and
achievement. As noted above, these biomarkers reflect exposures during prenatal, postnatal and childhood
lifestages (tooth Pb concentration may reflect exposure up to the time a tooth is shed depending on the
layer analyzed [see Section 2.3.4.1]). Most studies examined representative populations and had moderate
to high follow-up participation with no indication of selective participation among children with higher
BLLs and lower cognitive function. Associations between BLL and cognitive function decrements were
found with adjustment for several potential confounding factors, most commonly SES, parental IQ,
parental education, and parental caregiving quality. In children aged 4-11 years, associations were found
with prenatal (i.e., maternal or cord BLLs), early childhood, childhood average, and concurrent BLLs in
populations with mean or group BLLs in the range of 2-8 (ig/dL. No critical lifestage or specific duration
of Pb exposure within childhood was uniquely associated with cognitive function decrements based on
consideration of evidence from epidemiologic and toxicological studies. Several epidemiologic studies
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found a supralinear concentration-response (C-R) relationship (i.e., larger incremental effect at lower
BLLs). A threshold for cognitive function decrements was not discernable from the available evidence
(i.e., examination of early childhood blood Pb or concurrent [with peak <10 |ig/dL| blood Pb in the range
of <1 to 10 (ig/dL). Evidence in children was clearly supported by observations of Pb-induced
impairments in learning and memory in juvenile animals. Several studies in animals indicated learning
impairments with prenatal, lactational, post-lactational, and lifetime (with or without prenatal) Pb
exposures that resulted in BLLs of 10-25 (ig/dL. Biological plausibility for Pb-associated cognitive
function decrements was supported by observations of Pb-induced impairments in neurogenesis,
synaptogenesis and synaptic pruning, LTP, and neurotransmitter function in the hippocampus, prefrontal
cortex, and nucleus accumbens.
The structure of the current assessment of cognitive effects in children is similar to that in the
2013 Pb ISA. Although the above measures of cognitive function are interrelated, the evidence for each of
these categories of outcomes (i.e., FSIQ, Bayley Scales of Infant Development [BSID],
neuropsychological tests of learning, memory, and executive function, and academic performance) was
assessed separately, to the extent possible, in the order of strength of evidence. Studies assessing
cognitive function of school-age children using instruments that measure FSIQ are described in Section
3.5.1.1, and studies assessing cognitive development in infants using the BSID and other instruments are
described in Section 3.5.1.2. Studies examining the associations of Pb exposure with outcomes on
neuropsychological tests of learning and memory and executive function in children as well as analogous
endpoints in animals are discussed in Sections 3.5.1.3 and 3.5.1.4, respectively. These sections are
followed by a discussion of studies that examine the association of Pb exposure with academic
achievement and performance in Section 3.5.1.5. The final sections discuss issues relevant for the
interpretation of the evidence base (Section 3.5.1.6) and the summary and causality determination
(Section 3.5.1.7).
Because the conclusion from the 2013 Pb ISA was "causal," the PECOS statement for studies of
cognitive effects in children (see Section 3.2) was refined to emphasize recent studies that examined
lower BLLs more similar to those of current U.S. children (i.e., <5 (ig/dL). Details of these studies are
extracted into the evidence inventories (Section 3.7, Table 3-2E [FSIQ], Table 3-3E [Infant
Development], Table 3-4E [Learning, Memory, and Executive Function], and Table 3-5E [Academic
Achievement and Performance]). Studies with central tendency blood Pb concentrations that exceed 5
(ig/dL are extracted into Table 3-6E of Section 3.7. In addition to refining the PECOS statement to focus
on lower exposure levels, studies of younger children whose BLLs were less influenced by higher past Pb
exposures are considered particularly informative. Controls for important potential confounders identified
in the 2013 Pb ISA such as SES, parental education, quality of parental caregiving (often measured as the
HOME score), nutritional status, and birth weight in studies of postnatal Pb exposure were considered
attributes of high-quality studies (see section 4.3.13 of the 2013 Pb ISA (U.S. EPA. 2013)). A summary
of the recent evidence, which is interpreted in the context of the entire body of evidence, is provided in
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the subsequent sections. Overall, recent studies add to the evidence generally supporting the findings from
the 2013 Pb ISA.
3.5.1.1 Full-Scale IQ in Children
A large number of studies evaluated in the 2013 Pb ISA found a consistent pattern of associations
between higher BLL and lower FSIQ in children aged 4-17 years (see Figure 4-2 and Table 4-3 (U.S.
EPA. 2013)). FSIQ has strong psychometric properties (i.e., reliability, consistency, validity), is among
the most rigorously standardized cognitive function measures, is relatively stable in school-age children,
and has been demonstrated to be predictive of educational achievement and life success. The strongest
evidence was provided by prospective studies with analyses of the association of blood Pb levels
measured in early childhood or tooth Pb level that generally reflect the early childhood Pb exposure (i.e.,
prospective studies where Pb exposure preceded the assessment of FSIQ). These prospective studies
typically considered potential confounding by maternal IQ and education, SES, birth weight, smoking
exposure, parental caregiving quality, and in a few cases, other birth outcomes and nutritional factors.
Associations were found in diverse populations (e.g., Boston, MA; Cincinnati, OH; Rochester, NY;
Cleveland, OH; Mexico City, Mexico; Port Pirie, Australia; and Kosovo, formerly of Yugoslavia) in
studies that examined children recruited from prenatal clinics, hospital maternity departments, or schools.
Studies generally reported high follow-up participation, which was supported by evidence that selection
bias did not explain the associations observed. The few studies reporting weak or null associations (i.e.,
Cleveland, Sydney cohorts) were not stronger with respect to methodology or control for potential
confounding and did not weaken the far larger body of supporting evidence (U.S. EPA, 2013).
The blood Pb-FSIQ association in children was further substantiated by an international pooled
analysis of seven prospective cohorts (Lanphear et al„ 2019, 2005) as well as multiple meta-analyses that
combined results across various prospective and cross-sectional studies (Pocock et al„ 1994; Schwartz,
1994a; Needleman and Gatsonis, 1990). Schwartz (1994a) additionally demonstrated the robustness of
evidence to potential publication bias. The pooled analysis (Lanphear et al., 2019, 2005) examined several
BLL metrics and demonstrated that early childhood and concurrent childhood BLLs explained more
variation in FSIQ compared with the other blood Pb metrics, as indicated by the R-square values. The
coefficient for concurrent BLL had a smaller standard error (SE) than the coefficient for early childhood
BLL (Lanphear et al., 2019). Across studies, no clear indication that Pb exposure during one critical
lifestage or time period within childhood was uniquely or more strongly associated with FSIQ (see
Section 3.5.1.6.3). Blood Pb-associated FSIQ decrements at ages 4-17 years were found with concurrent,
prenatal (maternal or cord), early childhood (e.g., age 2 or 4 years), multiple-year average, or lifetime
average BLLs. Associations were also found with tooth Pb levels.
Key statistics associated with the international pooled analyses of seven cohort studies are
presented in Table 3-1. The C-R function was nonlinear, with a larger incremental effect of Pb on IQ at
lower blood Pb concentrations (Lanphear et al., 2019, 2005). The log-linear model coefficient (i.e., |3
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coefficient) for concurrent BLL was -2.65 (95% confidence interval [CI]: -3.69, -1.61) per unit change
in natural log transformed BLL. The linear association observed for a subset of 103 children with peak
BLLs <7.5 (mean concurrent BLL = 3.2 (ig/dL) was -2.53 (95% CI: -4.48, -0.58). Linear coefficients for
higher BLLs and using a peak BLL cutoff point of 10 (ig/dL are included in Table 3-1, as are other key
statistics, including R-square values for various models.
Table 3-1 Statistics associated with the international pooled analysis of data
from seven cohort studies
Main Finding from Analyses of the Pooled Dataset
Quantitative Result3
Log-linearb model coefficient for blood Pb metrics and IQ, adjusted for
site, HOME score, birth weight, maternal IQ, and maternal education
Early childhood: -2.21 (-3.38, -1.04)
Peak: -2.86 (-4.10, -1.61)
Lifetime average: -3.14 (-4.39, -1.88)
Concurrent: -2.65 (-3.69, —1.61 )c
IQ decrement over different concurrent blood Pb ranges based on the
log-linear model
2.4 to 30 |jg/dL: 6.7 IQ pts (4.1-9.3)
2.4 to 10 |jg/dL: 3.8 IQ pts (2.3-5.3)
10 to 20 |jg/dL: 1.8 IQ pts (1.1-2.6)
20 to 30 |jg/dL: 1.1 IQ pts (0.7-1.5)
Linear coefficient,11 sample size (n) and concurrent BLL measurements
(mean, minimum, 5th and 95th percentiles, and maximum) for subset
with peak BLLs <7.5 |jg/dL
-2.53 (-4.48, -0.58 )
n = 118e
(3.3, 0.9, 1.1, 6.7, 7.4 pg/dL)
Linear coefficient,11 sample size (n) and concurrent blood Pb
measurements (mean, minimum, 5th and 95th percentiles, and
maximum) for subset with peak BLLs >7.5 |jg/dL
-0.15 (-0.19, -0.11)
n = 1215
(13.0, 0.1, 3.7, 34.2, 71.7)
Linear coefficient,11 sample size (n) and concurrent blood Pb d
measurements (mean, minimum, 5th and 95th percentiles and
maximum) for subset with peak blood Pb <10 |jg/dL
-0.77 (-1.65, 0.12)
n = 258
(4.4, 0.1, 1.4, 8.0, 9.8)
Linear coefficient,11 sample size (n) and concurrent blood Pb
measurements (mean, minimum, 5th and 95th percentiles, and
maximum) for subset with peak BLLs >10 pg/dL)
-0.13 (-0.22, -0.04)
n = 1075
(14.0, 0.1,4.4, 35.5, 71.7)
Blood Pb metric with the largest R2 for the relationship with IQ in the
log-linear models
Early childhood R2: 0.6433 = largest
Peak R2: 0.6401
Lifetime average R2: 0.6411
Concurrent R2: 0.6414
BLL = blood lead level; HOME = Health Outcomes and Measures of the Environment; IQ = intelligence quotient; Pb = lead; pts =
points.
aResults reported by Lanphear et al. (20191 and/or Crump et al. (20131 and confirmed by Kirrane and Patel (20141.
bCoefficients are not standardized, i.e., coefficients indicate the decrement in full scale IQ per unit of natural log transformed blood
Pb. Standardized estimates (i.e., standardized to a 1 unit increase for the 10th—90th percentile interval of the biomarker level and
assumed to be linear within this interval) for this study are found in Evidence Inventory (Section 3.7, Table 3-2E).
°Slopes ranged from -2.36 to -2.94 in sensitivity analyses of concurrent blood Pb-IQ association, which omitted one cohort at a time.
dLinear coefficients are standardized to a 1 |jg/dL increase in blood Pb
eThe number of children from Boston cohort with peak BLLs <7.5 |jg/dL was 28 after errors were corrected (Lanphear et al.. 20191.
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Several studies that conducted analyses stratified by BLL provide additional support for the
findings of Lanphear et al. (2005) and Lanphear et al. (2019). These studies comprise a compelling body
of evidence demonstrating a nonlinear C-R function (i.e., larger decrement in cognitive function per unit
increase in blood Pb level in children in the lower range of the study population blood Pb distribution) for
the association between BLL and intelligence. This evidence is described in the 2013 Pb ISA (Section
4.3.12, Figure 4-15, and Table 4-16 of the U.S. EPA (2013)). Particularly compelling evidence was
provided by analyses that examined prenatal or early childhood blood Pb levels or considered peak blood
Pb levels in school-aged children (Schnaas et al., 2006; Bellinger and Needleman, 2003; Canfield et al.,
2003a). The geometric mean Pb level in Schnaas et al. (2006) was 7.8 (ig/dL and the mean blood Pb
levels in Bellinger and Needleman (2003) and Canfield et al. (2003a) were 3.8 (ig/dL and 3.3 (ig/dL,
respectively.
In a recent analysis, Crump et al. (2013) examined the shape of the C-R function for the pooled
data using an alternative modeling strategy. Rather than model the natural log of BLL as was done in the
original analysis (Lanphear et al., 2019, 2005), Crump et al. (2013) modeled the natural log (In) of blood
Pb + 1, which has the property of equaling zero when untransformed BLL equals zero. The authors
applied F-tests to nested models containing both In (BLL + 1) and non-transformed BLL and found that
the linear coefficient did not improve the prediction of the model, indicating that In (BLL + 1) was a
better predictor across the full range of the data (e.g., 2.5-33.2, as 5th to 95th percentile concurrent
BLLs). In addition, Crump et al. (2013) considered confounding by additional covariates, which were
defined as site-specific in their final models. Despite the aforementioned differences in modeling
approach, the Crump et al. (2013) analysis corroborated the findings of the original analysis, providing
strong evidence in support of the nonlinear C-R function and the causal association between Pb exposure
and cognitive effects in children. The coefficient for the Crump et al. (2013) log-linear association
between concurrent BLL and IQ was -3.32 (95% CI: -4.55, -2.08), somewhat larger than that reported
by Lanphear et al. (2019) (see Table 3-1).
Notably, the international pooled analysis (Lanphear et al„ 2019, 2005) included data from seven
longitudinal cohorts that were initiated before 1995. The median concurrent BLL was 9.7 (ig/dL (5th and
95th percentiles: 2.5-33.2 (ig/dL) with included studies reporting limits of detection of 1 (ig/dL. Several
other longitudinal and cross-sectional studies included in the 2013 Pb ISA, however, conducted analyses
of children with mean BLLs <5 (ig/dL, collectively providing strong evidence of an association between
Pb exposure and FSIQ at lower BLLs (Kim et al., 2009; Chiodo et al„ 2007; Surkan et al„ 2007;
Bellinger and Needleman, 2003; Canfield et al„ 2003a) (see Figure 3-4).
Van Landingham et al. (2020) extended the analyses of the international pooled data described
above that were first examined by Lanphear et al. (2005), Lanphear et al. (2019) and later by Crump et al.
(2013). These authors identified "highly likely" confounders (HOME score, maternal education and
maternal IQ) using a combination of correlation analysis and backward selection with a criteria of p =
0.15 to retain covariates in the model. Van Landingham et al. (2020) also included interaction terms
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between BLL and each covariate. The results of their analyses were comparable to previous analyses that
did not consider interactions. Specifically, the beta coefficient for concurrent blood Pb in the loglinear
(blood Pb level +1) model with interaction terms was -4.945. Coefficients for the interactions of maternal
IQ, maternal education and HOME score with Pb were negligible (i.e., -0.0003, -0.0051, and 0.0437,
respectively). The authors also fit predictive models and calculated the IQ decrement for various
increments of BLL, and across levels of each of the covariates. The interpretation of the covariates and
the predicted IQ decrements at various combinations of BLL decrement and covariate levels that are
presented in the paper, however, potentially conflate the interpretation of direct versus total effect
estimates and may not inform causality (i.e., the table 2 fallacy) (Westreich and Greenland. 2013). As
noted previously, there is a lack of data in the international pooled dataset for children with BLLs <1
microgram per dL and the data is sparse below <5 (ig/dL; hence, uncertainty remains regarding the C-R
function associated with this dataset in the range where the data are sparse and this uncertainty is not
addressed by Van Landingham et al. (2020).
Several recent longitudinal studies add to the evidence informing the relationship between BLL
and IQ in children. Heterogeneity in the magnitude and direction of the associations, which was
potentially explained by race/ethnicity, sex, and modeling choices such as adjustment for other metals or
chemicals was present. This heterogeneity did not weaken the larger body of supporting evidence.
Overall, recent studies generally corroborated previous epidemiologic observations of associations
between Pb exposure and IQ in children with relatively low blood Pb concentrations (<5 (ig/dL) (see
Figure 3-4 and Evidence Inventory Table 3-2E).
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Study
Prospective Studies
fTatsuta et al. 2020
tTaylor et al 2017
tZhou et al. 2020
tLee et al. 2021
Bellinger et al. 2003
tlglesias etal. 2011
tLee et al. 2021
tLee et al. 2021
Tohuku district. Japan Prenatal
Prenatal
Multi-center. United Kingdom Prenatal
Prenatal
Jiangsu Province, China Prenatal
Prenatal
Prenatal
Prenatal
8 hospitals. S Korea
Boston area, MA
Northern Chile
8 hospitals, S Korea
8 hospitals, S Korea
Lanphear et al. 2005, 2019 International
Canfield et al. 2003 Rochester, NY
Kim et al. 2009 4 Cities, Korea
tTatsuta et al. 2020
tRuebner et al. 2019
tlglesias etal. 2011
Tohuku district, Japan
Multi-center U S
Northern Chile
Cross-sectional Studies
tDantzer et al. 2020
tMartin et al. 2021
Cincinnati. OH
East Liverpool, OH
tHong et al. 2015 5 regions, S Korea
tMenezes-Filho et al. 2018 Bahia. Brazil
tLucchini et al. 2012 Brescia. Italy
Mean
(pg/dL)
0.8
3.67
1.59
1.32
Boys
Girls
All
Boys
Girls
Early child: 2 yrs
Early child: 3-6 yrs 10.8
Early child: 4 yrs 1.41 (med)
Early child: 6 yrs 1 44(med)
Concurrent
Concurrent
Concurrent
Concurrent
Concurrent
Concurrent
Concurrent
Concurrent
Concurrent
Concurrent
Concurrent
Concurrent
Concurrent
3.2
3.3
1.7
1.2
3.5
0.57
1.13
1.8
1.64
1.71
Pre-site closure
<7.5 pg/dL
<10 pg/dL
Low Mn
High Mn
Boys
Girls
CKD patients
FSIQ Adjusted for:
Age (yrs)
Hg, child Pb
Mn, Cd
Mn. Cd
Mn, Cd
Mn, Hg. Cd
10
10-13
5-15
5-15
4-10
5
6-10
1-18
Post site closure 7-16
High Hair
Low Hair Mn
12
7-9
8-11
7-12
11-14
Hg. prenatal Pb *
-10.00
-8.00
-6.00 -4.00 -2.00 0.00 2.00
Beta values (95% CI) per 1 ug/dL increase in blood Pb
Note: Effect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 jjg/g increase in bone Pb. If the Pb biomarker is log-transformed, effect estimates are standardized to
the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval. Categorical effect
estimates are not standardized. The "adjusted for" column indicates covariates that are not typically considered in multivariate models. The exhaustive list of confounders for individual
studies is found Table 3-2E.
tStudies published since the 2013 Integrated Science Assessment for Lead.
Figure 3-4 Associations between blood Pb levels and full-scale intelligence quotient in children.
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Several recent prospective epidemiologic studies were also conducted that examined the
associations of prenatal or postnatal BLLs and effect measure modification by sex. Taylor et al. (2017)
used data from the Avon Longitudinal Study of Parents and Children (ALSPAC) to study the association
of maternal BLL and child IQ at ages 4 and 8 (i.e., Wechsler Preschool and Primary Scales of Intelligence
[WPPSI] and Wechsler Intelligence Scale for Children [WISC]-III, respectively). Little evidence of an
association with IQ was observed. The change in score on the WPPSI associated with maternal BLL was
-0.32 (95% CI: -1.32. 0.68) per (ig/dL and the change in score on the WISC-III was 0.26 (95% CI:
-0.21, 0.73) per (ig/dL. Sex-stratified analyses, however, indicated an association between maternal BLL
and IQ decrement in boys (-0.29 [95% CI: -1.02, 0.44]), but not in girls, among whom positive
associations were observed (0.73 [95% CI: 0.39, 1.33]). Models were adjusted for covariates including
maternal education and indicators of SES. The mean prenatal and postnatal BLLs were 3.67 (ig/dL and
4.22 (ig/dL, respectively.
Tatsuta et al. (2020) examined the association of both cord BLL and postnatal BLL (at age 12)
with IQ (WISC-IV) among boys and girls enrolled in the Tohoku Study of Child Development, a
prospective birth cohort. In addition, the Boston Naming Test (BNT) was administered to assess language
abilities. This study found decrements in FSIQ score in association with postnatal BLL [|3 = —9.88 (95%
CI: -18.98, -0.78] among boys and a less precise association with IQ decrement that included the null
value among girls [|3 = -4.41 (95% CI: -15.94, 7.13]). Confounders considered in the analysis included
maternal IQ, parental SES (i.e., income) and Hg concentration in cord blood. Prenatal BLL was
associated with a relatively weak and imprecise decrease in FSIQ score among boys [|3 = -3.68 (95% CI:
-10.71, 3.35], The association of prenatal BLL with FSIQ was slightly positive but with the CI including
the null value among girls (|3 = 1.46 (95% CI: (-2.91, 5.83)]. Lower BNT scores (with cues) were
associated with both prenatal and postnatal BLL among boys. The associations of pre- and postnatal
BLLs with BNT were relatively weak or null in girls. The median postnatal BLL was 0.7 (ig/dL and the
median cord BLL was 0.8 (ig/dL in this study.
Desrochers-Couture et al. (2018) studied the association between cord, maternal, and childhood
(3-4 years old) BLLs with cognitive function (WPPSI-III at 3-4 years of age) among children enrolled in
the Maternal-Infant Research on Environmental Chemicals (MIREC) Study. The analysis included
mothers who participated in MIREC Chemical Study Plus (n = 610), which was conducted when the child
reached the age of 3-4 years old. The cohort from which the study participants were drawn comprised
Canadian preschoolers from mainly middle- to upper-middle SES families with low Pb exposure. The
geometric mean concurrent blood Pb concentration was 0.70 (ig/dL, and the geometric mean cord blood
Pb concentration was 0.76 (ig/d. Outcomes included FSIQ, verbal IQ, performance IQ, and a general
language composite. The authors report standardized regression coefficients for the associations of cord
blood Pb level and child blood Pb level with FSIQ. An association between cord blood Pb level and FSIQ
was observed [|3 = -0.07 (95% CI: -0.143, 0.003)], while the association of childhood concurrent BLLs
with FSIQ effectively null (|3 = 0.014 [95% CI: -0.071, 0.098], These results describe a SD change in
FSIQ (i.e., 13.5 points) per SD change in Log2-transformed blood Pb level (SD of Log2-transformed
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BLLs not reported; results not depicted in Figure 3-4). The cord blood model adjusted for child age, sex,
maternal education, evaluation site, and cord blood Hg, while the postnatal model adjusted for child age,
sex, evaluation site, marital status, income, HOME score, Parenting Stress Index, and cord BLL. No
pattern of Pb-associated cognitive function decrements emerged with the verbal or performance
components of IQ or with the general language composite. An association (non-standardized beta
coefficients) was observed between cord BLL and performance IQ in boys (|3 = -5.69 [95% CI: -9.97,
-1.41] but not in girls (|3 = 0.29 [95% CI: -3.79, 4.36]). The study adjusted for several important
confounders, including SES. Although maternal education was included in the model, maternal IQ, which
is a strong predictor of child IQ, was not accounted for in the analysis.
Zhou et al. (2020b) conducted a study to examine the association of cord blood concentrations of
trace elements, including Pb, manganese (Mn), and cadmium (Cd), with FSIQ, verbal IQ, and
performance IQ components among children from an agricultural region in China who were enrolled in a
prospective birth cohort. In models including each of these elements, cord BLL was not associated with
FSIQ (|3 = 0.67 [95% CI: -0.51, 1.85]). A similar lack of association was observed with performance and
verbal IQ and in sex-stratified analyses. Models were adjusted for covariates including maternal education
and family income. Each of the trace elements were included in the models but interactions between the
elements were not examined. The mean cord BLL among the children was 1.59 (ig/dL. Liu et al. (2015)
developed a predictive model to examine the association of Pb, Cd, and Hg in serum with FSIQ at age 5,
dropping variables based on the variance inflation factor (VIF >10). The final model for FSIQ (i.e.,
WPPSI) did not include cord serum Pb level; thus, no results pertaining to the association of Pb
concentration in serum with FSIQ were presented. Wang et al. (2022) investigated associations between
cord and concurrent venous blood concentrations of Pb, selenium (Se), As, Cu, Mn, and Cr and FSIQ
among children (6-8 years old) born in a hospital in Wujiang, Jiangsu Province. The geometric mean
concentrations of cord and venous blood Pb were 2.83 (ig/dL and 3.30 (ig/dL, respectively. Cord blood Pb
was weakly associated with reduced performance IQ (|3 = -0.11 [95% CI: -0.25, 0.03]) in boys, and
concurrent venous blood Pb was associated with reduced verbal IQ in girls (|3 = -0.49 [95% CI: -0.86,
-0.12]).
Lee et al. (2021) studied the association of Pb and other metals (i.e., Cd, Hg, and Mn) among
mother and infant pairs from eight hospitals in South Korea. In multivariable models including each of
the metals as well as covariates, imprecise negative associations of prenatal Pb exposure (|3 = -1.20 [95%
CI: -4.87, 2.01]), child BLL at age 4 (|3 = -1.83 [95% CI: -4.66, 1.01]), and child BLL at age 6 (|3 =
-2.61 [95% CI: -5.62, 0.40]) with FSIQ were observed. In another study of exposure to multiple trace
metals conducted in Wujiang, China, imprecise associations of maternal cord and early childhood venous
blood were observed with FSIQ (e.g., -4.77 [95% CI: -14.34, 4.79] comparing the upper quartile of
venous child BLL with the reference quartile) (Wang et al.. 2022). The geometric mean blood Pb
concentration was 2.30 (ig/dL (interquartile range [IQR]: 1.83-3.30 (.ig/dL). and the study included 113
children.
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The recent body of evidence also includes a which evaluated the association of BLL in early
childhood (3-6 years) and BLL later in childhood (10-13 years) when IQ was also measured. The BLLs
that were measured later in childhood corresponded to the period when a major source of Pb exposure
was eliminated (i.e., following the closure of a Pb storage facility) (Iglesias et al.. 2011). The early
childhood mean BLL was 10.8 (ig/dL, and concurrent mean BLL was 3.5 (ig/dL. This study found an
FSIQ decrement associated with concurrent BLL (|3 = -0.94 [95% CI: (-1.77, -0.11)]) and a weaker, less
precise association with early childhood BLL (|3 = -0.14 [95% CI: -0.45, 0.16]). Verbal IQ was more
strongly associated with concurrent BLL than performance IQ. These associations were adjusted for
important confounders including maternal IQ and education, HOME score, and SES; however,
participation was moderately low with approximately 43 percent of the children with early childhood
BLLs participating in the IQ assessment.
Braun et al. (2018) conducted a study to determine whether residential exposure interventions
would reduce BLL in children and further result in improvements in the IQ score assessed using the
WPPSI at 5 to 8 years of age. Eligible women were randomly assigned to either a Pb exposure reduction
or injury prevention group. The geometric mean BLLs for children from 1 to 8 years of age was 1.6
(ig/dL in the Pb exposure intervention group and 1.7 (ig/dL in the control group. Dust Pb loadings were
lower following the intervention, but no differences in BLL (i.e., risk of having blood Pb concentration
>2.5 or 5 (ig/dL) were observed among the children. The effect of the intervention on BLL differed
depending on race/ethnicity, however. Specifically, the relative risk (RR) of having an elevated BLL
(>2.5 (ig/dL) indicated a protective effect of the intervention among non-Hispanic black children who
received the Pb intervention (RR: 0.6 [95% CI, 0.4-1.0]), but not among non-Hispanic white children
who received the Pb intervention (RR: 1.0; 95% CI, 0.5-1.9; race/ethnicity x intervention p-value = 0.06).
No improvement in FSIQ was observed among children who received the Pb intervention, nor did
race/ethnicity modify the effect of the intervention on FSIQ.
Several cross-sectional analyses were also conducted. Dantzer et al. (2020) analyzed data drawn
from the Cincinnati Childhood Allergy and Air Pollution Study (CCAAPS), a longitudinal study that
followed children beginning at age 1 and included their caregivers. Children were assessed using the
WISC-IV at their age-12 study visit. BLL, toenail Pb concentration, and information on covariates were
also ascertained at the age-12 visit. A strong but imprecise association between BLL at age 12,
concurrently ascertained IQ -10.87 [95% CI: -16.89, -4.85]), and a relatively smaller association with
toenail Pb concentration (-1.70 [95% CI: -4.27, -0.86]) were observed after adjustment for caregiver IQ,
SES, BMI (sex was considered as a potential confounder). Toenail Pb concentration reflects blood Pb
concentration approximately months to a year before concurrent BLL due to the time it takes toenails to
grow. The concurrent BLL in this study was 0.57 (ig/dL.
Martin et al. (2021) found interactions between blood Pb and blood Mn level with IQ decrement
among children (n = 57-62 depending on the analysis) enrolled in the Communities Actively Researching
Exposure Study (CARES) cohort in East Liverpool, Ohio. BLL was measured and FSIQ ascertained at
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the first clinic visit, which occurred when the child was between 7 and 9 years of age. Stronger
associations between BLL and FSIQ were observed with increasing Mn concentrations in hair and
toenails. For example, the association of blood Pb with FSIQ ranged from 1.69 (95% CI: -3.04, 6.41)
when In hair Mn equaled 5 ng/g to -10.60 (95% CI: -17.17, -4.02) when In hair Mn equaled 7 ng/g. In
contrast, relatively imprecise associations between BLL and FSIQ at varying levels of blood Mn were
observed. The mean BLL among children in this study was 1.13 (ig/dL (range: 0.30-6.64). Havnes et al.
(2015) examined the association of BLL and FSIQ among the same cohort of children. The primary
objective of this study was to examine the effect of Mn on child intelligence. Associations between BLL
and FSIQ were not reported, although a 1 (ig/dL increase in blood Pb was associated with lower
processing speed (|3 = -3.53 [95% CI: -6.95, -0.12]).
Ruebner et al. (2019) evaluated the association between BLLs and FSIQ among children with
chronic kidney disease (CKD). FSIQ was assessed using several instruments depending on the child's age
(i.e., Mullen Scales of Early Learning [age 12-29 months], WPPSI [30 months-5 years], and Wechsler
Abbreviated Scale of Intelligence [WASI; 6-18 years]). Concurrent BLL assessment was associated with
FSIQ decrement (|3 = -2.1 [95% CI: -3.9, -0.2]). Covariates considered as potential confounders
included race, poverty, maternal education, and factors related to CKD (i.e., CKD stage, duration,
glomerular versus non-glomerular diagnosis, hypertension, proteinuria, and anemia). The median BLL in
this study was 1.2 (ig/dL.
Hong et al. (2015) found an association between blood Pb concentration and lower FSIQ (-2.12
[95% CI: -3.79, -0.45] in across-sectional analysis of Korean school children from 8 to 11 years old.
This association persisted in models adjusted for paternal education and income, ADHD rating scale
score, Mn, and Hg (-1.95 [95% CI: -3.61, -0.29] per 10-fold increase). The mean BLL in this study was
1.80 (ig/dL.
Menezes-Filho et al. (2018) examined the association of concurrent BLL with intelligence
(WASI) among children from 7 to 12 years old. Mn in hair and toenails was also measured and the
interaction between metals evaluated. Child IQ was associated with BLL in this study (-2.78 [95% CI:
-4.66, -0.89]) in adjusted models. The mean BLL of children in this study was 1.64 (ig/dL. The effect of
BLL on child IQ was greater among children with higher toenail Mn concentrations.
Lucchini et al. (2012) conducted a cross-sectional analysis of children between the ages of 11 and
14 to examine the relationship between concurrent BLL and FSIQ as well as the potential interactions
with Mn and the aminolevulinic acid dehydratase (ALAD) genotype. A decrement in FSIQ score was
observed in association with BLL after adjustment for covariates including SES and maternal education
(|3 = -2.24 [95% CI: -4.10, -0.37]). No interaction with Mn or ALAD was found. The mean BLL was
1.71 (ig/dL in this study.
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3.5.1.1.1 Summary
A large number of studies evaluated in the 2013 Pb ISA found a consistent pattern of associations
between higher BLL and lower FSIQ in children aged 4-17 years (U.S. EPA, 2013). Multiple recent
longitudinal studies add to the evidence informing the relationship between BLL and IQ in children.
Heterogeneity in the magnitude and direction of the associations was present across studies. In a study of
Canadian preschool children with low blood Pb levels, an association between cord blood Pb level and
FSIQ was observed, while the association of childhood concurrent BLLs with FSIQ effectively null
(Desrochers-Couture et al„ 2018). In addition, associations were observed in boys but not in girls in
several studies (Tatsuta et al., 2020; Taylor et al., 2017). There was some indication that the heterogeneity
across studies could be explained by modeling choices such as confounder adjustment for other metals.
For example, cross-sectional analyses found evidence that exposure to Mn may modify the association
between Pb exposure and IQ in some populations (Martin et al., 2021; Menezes-Filho et al„ 2018).
However, studies that adjusted for multiple metals (e.g., Mn, Hg, Cd, and Pb) in regression models,
without examining the interaction between metals, found little evidence of an association between cord or
postnatal BLL and IQ (Zhou et al., 2020b; Liu et al., 2015), imprecise associations only in boys (Tatsuta
et al., 2020), or large IQ decrements after adjustment for Mn, Hg, and ADHD rating score (Hong et al.,
2015). Overall, recent studies generally corroborated the epidemiologic observations of associations
between Pb exposure and IQ in children with relatively low blood Pb concentrations (<5 (ig/dL) among
some populations of children (see Figure 3-4 and Evidence Inventory Table 3-2E). Consistent with
findings from the 2013 Pb ISA, individual studies continue to report associations of FSIQ with prenatal
BLL (maternal and cord blood Pb) and postnatal BLLs measured at various childhood lifestages. The
heterogeneity in the observations across studies did not weaken the larger body of evidence supporting
the association of Pb exposure with cognitive effects in children at BLLs <5 (ig/dL.
3.5.1.2 Infant Development
The BSID is an assessment instrument that was developed to identify children with
developmental delays. The early versions (e.g., (Bavlev. 1969)) have been expanded and refined, with
subsequent versions incorporating three domains of development (i.e., cognitive, language, and motor),
and parent-reported subtests that reflect social, emotional, and adaptive behaviors (Albers and Grieve.
2007). The current version of the BSID, the BSID-IV, retains the same number of domains but includes
fewer questions within each domain and requires less time to complete (Balasundaram and Avulakunta.
2021).
This section focuses on the Mental Development Index (MDI) and the cognitive and language
scales of later versions of the BSID. The MDI and cognitive/language scales are reliable indicators of the
current development and cognitive function of infants, integrating cognitive skills such as sensory and
perceptual acuities, discriminations, and response; acquisition of object constancy; memory learning and
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problem-solving; vocalization and beginning of verbal communication; and basis of abstract thinking
(McCall et al., 1972). However, the MDI test is not an intelligence test, and MDI scores, particularly
before ages 2-3 years, are not necessarily strongly correlated with later measurements of FSIQ in children
with normal development (U.S. EPA, 2013).
In the review of the MDI evidence in the 2013 Pb ISA, emphasis was placed on results from
examinations at ages 2-3 years, which incorporate test items more similar to those in school-age IQ tests.
Most of the prospective studies reviewed in previous ISAs (U.S. EPA, 2013, 2006) found associations of
higher prenatal (cord and maternal BLL), earlier infancy, and concurrent BLL with lower MDI scores in
children aged 2 to 3 years (see Table 4-4 of the 2013 Pb ISA). These blood Pb-associated decrements in
MDI were observed in populations with mean BLLs of 1.3 to 7.1 (ig/dL. Studies typically recruited
participants before or at birth without consideration of Pb exposure or maternal IQ and reported high to
moderate follow-up participation as well as nondifferential loss-to-follow-up. Most studies adjusted for
birth outcomes, maternal IQ, and education. Cord BLLs were associated with MDI, with additional
adjustment for SES and HOME score in the Boston cohort (Bellinger et al„ 1987) and for HOME score in
the Yugoslavia cohort (Wasserman et al., 1992). Some studies found a stronger association of MDI with
prenatal BLLs than child postnatal BLLs (Hu et al., 2006; Gomaaet al„ 2002; Bellinger et al„ 1987).
Among the studies assessed in the 2013 Pb ISA, several included children with mean BLLs less
than 5 (ig/dL (Henn et al., 2012; Jedrychowski et al„ 2009b; Hu et al„ 2006; Bellinger et al„ 1987).
Recent longitudinal epidemiologic studies of populations or including groups with maternal, cord, or
postnatal mean BLLs less than 5 (ig/dL add to the overall body of evidence (see Section 3.7, Table 3-3E).
These studies are presented in Figure 3-5.
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Study
Location
Blood Pb
Mean
Age at Outcome
Strata
Prospective Studies
(pg/dL)
(months)
Bellinger et al. 1987
Boston, MA
Prenatal (cord)
6.5
24
Pb 6-7 vs. *10 pg/dL (ref mean: 14.6)*
Prenatal (cord)
1.8
24
Pb <3 vs. *10 pg/dL (ref mean: 14.6) *
Jedrychowski et al. 2009
Krakow, Poland
Prenatal (cord)
1.23 (med)
24
Prenatal (cord)
1.23 (med)
36
Hu et al. 2006
Mexico City, Mexico
Prenatal (T1)
7.07
24
Prenatal (T3)
6.86
24
Prenatal (cord)
6.2
24
TKimet al.2013
3 Cities, S Korea
Prenatal (early pregnancy)
1.4 (GM)
6
Prenatal (late pregnancy)
1.3 (GM)
6
Prenatal (early pregnancy)
1.4 (GM)
6
Cd <1.47 pg/L
Prenatal (late pregnancy)
1.3 (GM)
6
Cd >1.47 pg/L
Prenatal (early pregnancy)
1.4 (GM)
6
Cd<1.51 pg/L
Prenatal (late pregnancy)
1.3 (GM)
6
Cd>1.51pg/L *
fValeri et al. 2017
2 Districts. Bangladesh
Prenatal (cord)
1.8
20-40
Pabna
Prenatal (cord)
6
20-40
S i raj di khan
Hu etal. 2006
Mexico City, Mexico
Concurrent
4.79
24
Claus Henn et al. 2012
Mexico City, Mexico
Concurrent (12 months)
5.1
12-36
Concurrent (24 months)
5
12-36
Concurrent (12 months)
5.1
12-36
Mn <2 pg/dL
Concurrent (24 months)
5
12-36
Mn <2 pg/dL
1 1 1
-2.00 0.00 2.00
Beta values (95% CI) per 1 ug/dL increase in blood Pb
Note: Effect estimates are standardized to a 1 [jg/dL increase in blood Pb or a 10 fjg/g increase in bone Pb. If the Pb biomarker is log-transformed, effect estimates are standardized to
the specified unit increase for the 10th -90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval. Categorical effect
estimates are not standardized.
"("Studies published since the 2013 Integrated Science Assessment for Lead.
Figure 3-5 Associations between biomarkers of Pb exposure and Bayley Score of Infant Development
Mental Development Index.
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Several studies were conducted using data from Mexico City birth cohorts that enrolled low and
middle-income women seeking prenatal care at maternity hospitals belonging to the Mexican Institute of
Social Security (Y Ortiz et al.. 2017; Henn et al.. 2012; Hu et al.. 2006). Hu et al. (2006) and Sanchez et
al. (2011) were designed to elucidate the time window during pregnancy when the effect of Pb exposure
on neurodevelopment is most pronounced and are discussed in Section 3.5.1.6.3 and included in Table
3-6E, which includes studies with central tendency BLLs >5 (ig/dL. Y Ortiz et al. (2017) examined the
modification of the Pb-neurodevelopment association by prenatal stress using the Crisis in Family
Systems-Revised (CRISYS-R) questionnaire, which assesses negative life events across several domains
(i.e., financial, legal, career, relationships, community and home violence, medical problems, other home
issues, discrimination or prejudice, and difficulty with authority). Using structural equation models, this
study found that 3rd trimester maternal BLL (|3 = -6.60 [95% CI: -13.49, 0.29] per unit of log-
transformed BLL) and the quadratic term for stress (|3 = -0.23 [95% CI: -0.45, -0.01] per unit of log-
transformed BLL) were associated with lower scores on the cognitive component of the BSID. A weak
more than multiplicative interaction between 3rd trimester maternal BLL and stress was also observed (|3
= 1.02 [95% CI: -0.78, 2.82]). Approximately 67% of the mother-infant pairs had complete information
for covariates, which included maternal education, IQ, and HOME score. Henn et al. (2012) studied the
interaction between postnatal blood Mn and Pb levels (age 12 and 24 months) and MDI score at five
different time points between 12 and 36 months of age among the Mexico City mother-infant pairs. The
coefficients for the association between BLL at 12 and 24 months with MDI score were -0.07 (95% CI: -
0.39, 0.25) and-0.08 (95% CI: -0.46, 0.30), respectively. Interactions between the highest quintile of Mn
and continuous BLL at 12 months were observed (|3 = -1.27 [95% CI: -2.18, -0.37]). The model was
adjusted for covariates including hemoglobin, maternal IQ, and maternal education.
Kim et al. (2013b. 2013c) studied the combined effect of prenatal exposure to Pb and Cd on
infant cognitive development at 6 months of age among participants in the Mothers" and Children's
Environmental Health (MOCEH) study, which enrolled infant-mother pairs from maternity clinics in
three Korean cities. Higher maternal BLL in late pregnancy was associated with lower MDI scores (|3 =
-1.74 [95% CI: -3.37, -0.12]), while maternal BLL in early pregnancy was not (|3 = 0.02 [95% CI:
-1.20, 1.24] per (.ig/dL). This association was found after adjustment for Cd and other covariates
including maternal education and SES. A larger decrement in MDI was associated with late pregnancy
maternal BLL among those with Cd levels above the median (|3 = -3.20 [95% CI: -5.35, -1.06])
compared with the decrement observed among those with Cd levels below the median (|3 = -0.29 [95%
CI: -2.88, 2.30]). Further, an increase in MDI was observed in association with early pregnancy maternal
BLL among those with Cd levels below the median (|3 = 2.44 [95% CI: 0.04, 4.83]), indicating the
potential for random error, differential confounding, or other forms of bias to influence findings. In
another study of mother-infant pairs in Korea, Kim et al. (2018b) evaluated the associations between MDI
and various chemicals and metals, including Pb, in perinatal maternal whole blood and umbilical cord
blood. The median maternal and cord blood Pb concentrations were 2.7 (ig/dL and 1.2 (ig/dL,
respectively. Associations of blood Pb concentrations and MDI were assessed but not reported because
they lacked statistical significance.
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Valeri et al. (2017) examined the combined effect of cord blood concentrations of Pb, arsenic
(As), and Mn with cognitive and languages scores on the BSID. This study enrolled infant-mother pairs
from two birth cohorts in Bangladesh, which differed substantially regarding metal profiles and maternal
characteristics including maternal education. This study presents results from multiple regression modes
and also applied Bayesian kernel machine regression (BKMR) in a prospective analysis that considered
covariates including maternal IQ, education, and HOME score. A weak association between increasing
cord Pb level and decreasing cognitive score was observed in the group with lower Mn and As
concentrations in cord blood (|3 = -0.01 [95% CI: -0.02, 0.00]) but not in the group with higher
concentrations of these metals (|3 = 0.01 [95% CI: -0.05, 0.07]).
Koshv et al. (2020) analyzed data from a birth cohort following children living in a slum in
Vellore, India. Blood Pb concentration at 15 and 24 months was averaged to determine the association
with raw cognition score on the BSID at age 2 (|3 = -0.2 [95% CI: -0.2, -0.03]). These results were
adjusted for covariates including SES, maternal IQ, and iron level. In another study, Shekhawat et al.
(2021) obtained cord blood Pb data and BSID-III scores at 6.5 months on average in a prospective cohort
study of mother-child pairs in western Rajasthan, India. The linear regression models showed no
significant associations of Pb levels and cognitive or language scores.
Paraiuli et al. (2015a) and Paraiuli et al. (2015b) assessed the association of cord BLLs with MDI
at 24 and 36 months of age, respectively, in a birth cohort of mother-child pairs recruited from a general
hospital in Bharatpur, Nepal. The median blood Pb concentration was 2.06 (ig/dL. Adjusting for in utero
Pb, As, and zinc (Zn) levels, HOME score, mother's age, parity, mother's education level, family income,
mother's body mass index (BMI) just before delivery, weight of the infant at birth and 24 months after
birth, gestational age, and infant age at the time of BSID-II assessment, no association was observed
between cord blood Pb and 24-month MDI (|3 = -4.21 [95% CI: -13.62, 5.20] per log-transformed BLL)
or 36-month MDI (|3 = 4.05 [95%CI: -3.21, 11.31] per log-transformed BLL).
Several recent studies assessed neurodevelopment using other validated instruments (Nozadi et
al.. 2021; Nvanza et al.. 2021; Zhou et al.. 2017; Vigeh et al.. 2014; Lin et al.. 2013). Zhou et al. (2017)
assessed 139 mother-child pairs from the Shanghai Stress Birth Cohort. Maternal whole blood and
maternal prenatal stress levels were assessed at 28-36 weeks of gestation, and the Gesell Developmental
Schedules (GDS) adapted for a Chinese population were administered to children at 24-36 months of age
in the study. This instrument measures development quotients (DQs) in five domains (gross motor, fine
motor, adaptive behavior, language, and social behavior) and has been validated for children 0-84 months
old. For this section on neurodevelopment, only the language domain is relevant. The Symptom
Checklist-90-Revised was used to produce a Global Severity Index (GSI) for evaluating overall maternal
emotional stress. After controlling for child sex, age, maternal age, gestational week, birth weight,
maternal education, and family monthly income, there was no association between prenatal maternal BLL
and child cognitive development. However, the authors observed interaction effects such that high
maternal stress appeared to exacerbate the effect of prenatal Pb exposure in several domains, including
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language (|3 = -33.82 [95% CI: -60.04, -7.59] per log-10 transformed unit of BLL), while low maternal
stress did not (|3 = -1.76 [95% CI: -13.03, 9.51] per log-10 transformed unit of BLL).
Vigeh et al. (2014) evaluated 174 children in Tehran, Iran up to 36 months postpartum in eight
developmental areas (social, self-help, gross motor, fine motor, expressive language, language
comprehension, letters, and numbers) using Harold Ireton's Early Child Development Inventory (ECDI).
Items for these areas were combined to generate a general development ECDI score, with higher scores
representing better development. This parent-reported measure is meant for use with children 15 months
to 6 years old and includes 60 age-discriminating items from the Minnesota Child Development
Inventory. To assess Pb exposure, three maternal whole blood samples and one umbilical cord blood
sample were collected from each mother-child pair in the first, second, and third trimesters and at
delivery, respectively. The authors observed increased odds (odds ratio [OR] = 1.74 [95% CI: 1.18, 2.57])
of a low ECDI score (<20% lower than expected for the children's age and sex) in the first trimester (BLL
= 4.15 (ig/dL), adjusting for hematocrit, maternal education, BMI, family income, gestational age, birth
weight, and first born.
Lin et al. (2013) measured Pb and other metals (i.e., Mn, As, and Hg) in cord blood samples from
230 mother-infant pairs from the Taiwan Birth Panel Study (TBPS) and assessed development in
cognition, language, motor, social, and self-care skills among 2-year-old children with the Comprehensive
Developmental Inventory for Infants and Toddlers (CDIIT), which has been standardized for children 3 to
71 months old. The CDIIT uses DQs, and a score of 100 represents normal development. After adjusting
for maternal age, maternal education, fish intake >2 times/week during pregnancy, infant gender,
environmental tobacco smoke during pregnancy and after delivery, and HOME Inventory score, the linear
regression models showed that highly Pb-exposed (>75th percentile: 1.65 (ig/dL) children had lower
cognitive DQs (|3 = -5.35 [95% CI: -9.64, -1.06]) compared with those in the low-exposure (<75th
percentile) group. The authors also observed an interaction with Mn such that children who were highly
exposed to both Mn and Pb had larger deficits in cognitive (|3 = —8.19 [95% CI: -14.40, -1.98]) and
language (|3 = -6.81 [95% CI: -12.16, -1.46]) DQs compared with those with low exposure to just one or
both of these metals.
Nozadi et al. (2021) collected blood samples from pregnant mothers at the 36-week visit or at
time of delivery and administered the Ages and Stages Questionnaire Inventory (ASQ:I) at 10-13 months
of age to evaluate neurodevelopment. Trained staff scored children on five 65-70 item developmental
domains: communication, gross motor, fine motor, problem-solving, and personal-social. A 1 (ig/dL
increase in prenatal blood Pb was associated with small, imprecise decreases in problem-solving (|3 =
-0.67 [95% CI: -1.54, 0.20]) scores.
Nvanza et al. (2021) collected dried blood spots from a finger prick to measure Pb (in addition to
Hg, Cd, and As) concentrations in pregnant mothers at 16-27 weeks of gestation from the Mining and
Health study in Northern Tanzania. The authors used the Malawi Developmental Assessment Tool
(MDAT) translated into Kiswahili to assess several functional domains, including social development, in
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children between 6 and 12 months old. MDAT has been validated for children 0-6 years old in rural sub-
Saharan Africa. Covariates in the Poisson regression model included maternal age, maternal education,
maternal and parental occupation, number of under-5-year-old siblings at home, family socioeconomic
wealth quintile, infant sex, infant age, birth weight, and height and weight at the time MDAT was
administered. Concentrations of Pb were low (median: 2.72 |ig/dL). and the German Environmental
Survey for Children reference level of 3.5 (ig/dL was used to dichotomize Pb exposure groups into low
and high exposure groups. The authors did not observe significant associations between high Pb exposure
and language impairment. However, children highly exposed to both Hg (>0.08 (ig/dL) and Pb were more
likely to have global neurodevelopmental impairment (prevalence ratio [PR =1.4 [95% CI: 0.9, 2.1]).
3.5.1.2.1 Summary
Most of the prospective studies reviewed in previous ISAs (U.S. EPA, 2013, 2006) found
associations of higher prenatal (cord and maternal BLL), earlier infancy, and concurrent BLL with lower
MDI score in children aged 2 to 3 years (see Table 4-4 of the 2013 Pb ISA). These blood Pb-associated
decrements in MDI were observed in populations with mean BLLs of 1.3 to 7.1 (ig/dL. Studies typically
recruited participants before or at birth without consideration of Pb exposure or maternal IQ and reported
high to moderate follow-up participation and nondifferential loss-to-follow-up. Recent studies continue to
support associations between Pb exposure (i.e., maternal (Y Ortiz et al., 2017; Vigeh et al„ 2014; Kim et
al„ 2013b, c), cord (Valeri et al„ 2017), and postnatal exposure (Lin et al., 2013)) and poorer performance
on tests of neurodevelopment among mothers and infants with mean BLLs <5 (ig/dL (see Figure 3-5).
Although Zhou et al. (2017) found no association overall, this study reported decrements in several
domains of the GDS among infants of mothers reporting high maternal stress. Similarly, Y Ortiz et al.
(2017) found some evidence of interaction between Pb exposure and maternal stress. Several studies
found interactions between Pb, Mn, or Mn and As (Valeri et al„ 2017; Lin et al., 2013; Henn et al., 2012)
or Cd exposure (Kim et al., 2013b, c). The direction of the interaction was not consistent across studies.
Overall, recent studies support findings from the 2013 Pb ISA and extend the evidence pertaining to
modification of the association between Pb exposure and infant neurodevelopment by maternal stress and
exposure to other metals.
3.5.1.3 Learning and Memory
The 2013 Pb ISA included many studies examining the associations of blood Pb levels with
neuropsychological tests of memory and learning. These domains of cognitive function are related to
intelligence, and several were evaluated in the subtests of FSIQ. Further, indices of memory and learning
are comparable to endpoints examined in experimental animal studies.
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3.5.1.3.1
Epidemiologic Studies of Learning and Memory in Children
The studies evaluated in the 2006 Pb AQCD and the 2013 Pb ISA did not clearly indicate
associations between higher BLL and poorer performance on neuropsychological tests of memory or
learning (i.e., acquisition of new information) in children 4-17 years of age (see Table 4-5 (U.S. EPA,
2013)). The studies used various tests (e.g., spatial span total errors on the Cambridge
Neuropsychological Test Automated Battery [CANTAB], digit span or learning factor score on the
WISC, Kaufman Assessment Battery for Children [K-ABC], memory score on the McCarthy Scale of
Children's Abilities, California Verbal Learning Test [CVLT], and working memory on the Wide Range
Assessment of Memory and Learning [WRAML]) to assess learning and memory, which may account for
some of the heterogeneity observed in the findings. Notably, evidence for both memory and learning from
prospective analyses of several established cohorts (i.e., Rochester, Boston, and Cincinnati) was mixed
(Canfield et al., 2004; Ris et al., 2004; Stiles and Bellinger, 1993; Bellinger et al., 1991; Dietrich et al.,
1991). These prospective studies examined blood Pb metrics including early childhood, lifetime average.
Cross-sectional studies included in the previous ISA, however, generally found associations between
higher concurrent BLLs and poorer learning and memory, including the large (n = 4,853) study of
children aged 5-16 years who participated in the National Health and Nutrition Examination Survey
(NHANES) III (Lanphear et al„ 2000). Associations of higher concurrent BLL and poorer memory in
children aged 5-16 years were also observed by Krieg et al. (2010) and Froehlich et al. (2007); however,
some studies reporting such associations had limited implications because they lacked consideration for
potential confounding (Counter et al„ 2008; Min et al., 2007). Several studies included in the 2013 Pb
ISA were conducted in populations with mean BLLs <5 (ig/dL (Krieg et al., 2010; Surkan et al„ 2007;
Lanphear et al., 2000) and reported associations between increasing concurrent childhood blood Pb
concentration and lower performance on tests of learning and memory.
A small number of recent studies examined the association of Pb exposure with children's
performance on neuropsychological tests of learning and memory (see Section 3.7, Table 3-4E). Several
such studies examined the association between Pb exposure and performance on tests of learning and
memory in models that adjusted for several important confounders plus co-exposure to other metals or
chemicals. Yorifuji et al. (2011) evaluated the association of cord BLL with several components of IQ at
age 7 and age 14 in a Faroese birth cohort also exposed to methyl mercury (MeHg). IQ components
including attention and working memory, language, visuospatial reasoning, and memory were assessed
using the WISC-R and the children's version of the CVLT. The association of cord BLL with
neuropsychological tests of cognition was reported without adjustment for cord Hg, with adjustment for
cord Hg, and with a term for the interaction of cord blood Pb and cord Hg concentration. Poorer
performance on the digit span components of the WISC-R, which measure short-term memory, were most
consistently observed in association with cord BLL. The results for associations with performance on
some of the tests indicated that the interaction between Pb and methyl Hg (MeHg) may be less than
additive (i.e., the associations of cord blood with the neuropsychological test outcomes were most
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discernable among children with hair Hg concentrations below 2.61 |ig/g and among those with the
lowest cord Hg concentrations (e.g., |3 = -0.27 (-0.42, -0.11) at age 14).
Another recent study by Tatsuta et al. (2014) also examined exposure to multiple chemicals
including Pb, PCBs, and MeHg. The outcome in this study was performance on the K-ABC at 42 months
of age. No associations with sequential processing speed score (-2.14 [95% CI: -12.80, 8.53]) or mental
processing score (-3.32 [95% CI: -12.41, 5.77]) were observed after adjustment for variables including
other chemicals, maternal IQ, and family income (associations per unit of log [base not reported]
transformed BLL). Similarly, Oppenheimer et al. (2022) examined the association of cord BLL with
working memory assessed using the WRAML among children (13-17 years old) living near a superfund
site and thus exposed to multiple metals. Regression models were adjusted for prenatal concentrations of
dichlorodiphenyldichloroethylene (DDE), hexachlorobenzene (HCB), PCBs, Pb, and Mn as well as other
important confounders including HOME score and maternal IQ. The associations between verbal working
memory, symbolic working memory and working memory index differences were 0.12 (95% CI: -0.20,
0.45), 0.09 (95% CI: -0.25, 0.42), and 0.59 (95% CI: -0.97, 2.15) respectively. The interaction of Pb
exposure and sex was examined but no statistical evidence of the interaction was observed.
Summary
The studies evaluated in the 2006 Pb AQCD and the 2013 Pb ISA did not clearly indicate
associations between higher BLL and poorer performance on neuropsychological tests of memory or
learning (U.S. EPA, 2013). A small number of recent studies of children with mean BLLs <5 (ig/dL add
to the evidence informing the association of Pb exposure with performance on tests of memory and
learning; however, the results from these recent studies do not enhance the consistency of the evidence as
a whole. Some of the available studies consider co-exposure to other chemicals and metals as confounders
(Tatsuta et al., 2014) although there is evidence that such co-exposures may interact with or modify the
association between Pb and the outcome (Yorifuji et al., 2011). The evidence regarding the effect of Pb
exposure on specific tests of learning and memory lacks consistency, overall.
3.5.1.3.2 Experimental Animal Studies of Learning and Memory
As described in the preceding sections, BLLs are consistently associated with decrements in FSIQ
in children but show variable associations with performance on tests of learning and memory. A
relationship between Pb exposure and cognitive function deficits is further supported by evidence for Pb-
induced impairments in memory and learning in animal models. Critical evidence for the association of
Pb with cognitive impairment comes from a series of studies describing the effects of lifetime Pb
exposure on nonhuman primates (Rice, 1992; Rice and Gilbert, 1990a; Rice, 1990; Rice and Karpinski.
1988). Cynomolgus monkeys {Macaca fascicidaris) were dosed continuously from birth and tested
repeatedly throughout their lifetime. While these exposures yielded BLLs beyond values considered
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relevant for the current assessment (>30 |ig/dL). they provide key evidence of Pb-induced cognitive
impairments in a translationally relevant species.
Learning and Memory - Morris Water Maze
In rodents, spatial learning and memory have been evaluated using several paradigms, including
the Morris water maze. Typically, the Morris water maze task is separated into two distinct phases.
During the training phase, spatial learning is assessed by measuring the time or distance required for a
rodent to swim to a submerged platform using visual cues beyond the maze (e.g., basic shapes). Slower
decreases in time to escape from the maze across training trials (i.e., escape latency) can be indicative of
impaired spatial learning. After the animals have learned the location of the hidden platform (confirmed
by steadily decreasing escape latencies across training trials), memory for the location of the platform is
assessed in the probe phase by removing the platform and measuring the time each animal spends in that
area of the maze. Decreased time spent or distance swam in the target zone can be indicative of a deficit
in spatial memory. Although performance in both phases is primarily a function of learning and memory,
other impairments, such as decreased motivation, motor deficits, or altered perceptual function may also
influence the results. The impact of these factors is difficult to completely characterize, but some studies
may include additional controls or tests (e.g., baseline swimming activity) to reduce this uncertainty.
Because this ISA focuses on low exposure levels which typically do not cause overt toxicity, the impact
of these factors is less likely to play a major role in the interpretation of these results.
The 2013 Pb ISA (U.S. EPA, 2013) reviewed the evidence suggesting that exposure to Pb
produced learning and memory impairments in laboratory rodents using the Morris water maze. Several
studies involved exposures to Pb of varying durations and across different developmental periods.
Significant impairments in both learning and memory were reported for developmental exposures
resulting in BLLs ranging from 23 to 70 (ig/dL. For example, Kuhlmann et al. (1997) compared the
effects of Pb exposure during various lifestages and reported impaired learning and long-term memory in
adult Long-Evans rats exposed during gestation and lactation (via maternal diet) or over a lifetime from
gestation through adulthood. Each of the exposure periods examined produced peak BLLs of 59 (ig/dL.
Exposure during adolescence only, which produced BLLs of 23 (ig/dL, did not affect memory. In contrast
with Kuhlmann, other studies reviewed in the previous ISA reported that postweaning Pb exposure (8
weeks via drinking water) in Sprague Dawley rats resulted in significant deficits in both learning and
memory using the Morris water maze (Fan et al., 2010; Fan et al„ 2009). Recent studies (see evidence
inventory Table 3-4T) provide consistent evidence for Pb-induced impairments in learning and memory
following developmental exposures with lower BLLs than covered in the previous ISA (<30 (.ig/dL).
Evidence reviewed in the previous ISA indicated that development (i.e., preconception, during
gestation, lactation) may be a critical window for Pb exposure to cause cognitive dysfunction later in life.
Several recent studies examined the effects of long-term Pb exposure that began during development and
continued into adulthood. In the study with the longest exposure duration that was relevant to this ISA,
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Ouvang et al. (2019) developmentally exposed Sprague Dawley rats to Pb (0.05% Pb acetate in maternal
drinking water) beginning on GD 0. After weaning, animals were maintained on drinking water
containing (0.01% Pb acetate) until PND 679 (701 total days of exposure). This exposure resulted in a
final mean BLL of 22 (ig/dL. When assessed immediately following the end of exposure, exposed
animals displayed both impaired learning and memory in the Morris water maze task (31% fewer
crossings in the target zone during the probe trial compared with controls). Zhu et al. (2019b) exposed
rats to Pb (0.5 g/L Pb acetate in maternal drinking water) for 387 days beginning at conception, which
resulted in a final mean BLL of 29 (ig/dL. In the Morris water maze, exposed animals showed
significantly increased escape latencies on the last day of training only, suggestive of slightly impaired
spatial learning. In the probe trial, exposed animals made 45% fewer crossings into the target zone than
controls, strongly suggestive of impaired spatial memory. In another long-term study, Zhou et al. (2020a)
developmentally exposed Sprague Dawley rats to Pb through maternal drinking water beginning at
conception and continuing through lactation. After weaning, animals were maintained on Pb in drinking
water (386 days). While this study examined several doses of Pb, only the lowest dose (0.5 g/L in water)
produced BLLs that were relevant to this ISA. Learning and memory were tested via the Morris water
maze during exposure at PND 21 (mean BLL 10 (ig/dL) and later immediately following the end of
exposure at PND 364 (mean BLL 15 (ig/dL). At both time points, Pb-exposed animals took significantly
more time to escape the maze during training and spent less time in the target zone during the probe trial,
indicative of impaired learning and memory. The effect of Pb was slightly more pronounced at the earlier
timepoint (number of crossings in target zone was 36% lower than that of controls at PND 21, compared
with a 26% difference at PND 364), which may be due to improvement on the task with age.
In a study by Tartaglione et al. (2020). male and female Wistar rats were developmentally and
lactationally exposed to Pb beginning 4 weeks prior to conception (GD -28) to PND 23 (50 mg/L in
maternal drinking water), resulting in a final BLL of 26 (ig/dL, and displayed increased escape latencies
compared with controls. This effect was not sex-specific. Exposed animals showed mild memory deficits
in the form of increased latency to target zone and increased distance to target zone relative to controls
(no effect observed on crossings or time spent in target zone). While effects on memory were not sex-
specific, the authors reported that Pb significantly decreased path efficiency (ratio of the shortest possible
path length to the observed path length) in females only, which may indicate a sex-specific effect of Pb on
the processes that govern spatial integration.
Xiao et al. (2014) compared the effects of Pb on two separate developmental windows in Wistar
rats: one beginning prior to conception (2 mM Pb in maternal drinking water from GD -21 to PND 21,57
days total) and the other beginning in adolescence (2 mM Pb in drinking water from PND 21 to 84, 63
days). The gestational exposure yielded a final BLL of 10 (ig/dL at PND 21, while the adolescent
exposure led to a final BLL of 4 (ig/dL. Animals from both exposure groups were tested in the Morris
Water Maze on PND 85. Exposed animals had significantly increased escape latencies and decreased
target zone time relative to control animals, indicating cognitive dysfunction. No difference was observed
between the exposure time frames, suggesting both may be similarly vulnerable to the cognitive effects of
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Pb assessed by the Morris water maze. Similarly, Barkur and Bairv (2015b) employed a study design that
examined multiple different time frames of exposure: pregestational, gestational, and combined gestation
and lactation. All but the combined gestation and lactation group yielded BLLs that were relevant to this
ISA. Pb exposure during each period of development had a significant negative effect on memory
compared with the control groups (learning data were not reported). The gestation and lactation groups
exhibited similar magnitudes of effects, with the pregestational group showing the smallest difference
compared with the control. This study suggests that Pb affects memory following developmental
exposure and that the periods of gestation and lactation may be more sensitive than the pregestational
period alone.
Wang et al. (2021a) exposed Sprague Dawley rats to 0.05 and 0.1% Pb in drinking water from the
beginning of gestation through the end of lactation. BLLs assessed on PND 21 resulted in BLLs of 24.9
and 30.4 (ig/dL for the two dose groups, respectively. On PND 21, Pb-exposed rats displayed
significantly increased escape latencies during acquisition, indicative of impaired learning. During the
probe trial, only animals in the highest Pb concentration had significantly fewer crossings in the target
zone. Betharia and Maher (2012) exposed Sprague Dawley rats starting at conception (10 |ig/mL in
maternal drinking water, GD 0 to PND 20) and assessed cognitive function via the Morris water maze at
two points: end of exposure (PND 21, BLLs of 0.98 (ig/dL) and later (PND 56, BLLs of 0.03 (.ig/dL). This
study reported the lowest BLLs for animals tested in the Morris water maze paradigm. In contrast to many
of the recent studies reviewed here, when assessed immediately following the end of exposure, no effects
on learning or memory were observed. At the later time point, exposed females displayed significantly
impaired memory relative to untreated controls. This minor discrepancy could be due to the lower dose
used in the study. It is also possible that repeated experience with the paradigm across two sessions could
"unmask" a subtle effect on learning and memory produced by low-level exposure to Pb, though data on
cognitive function in animals with BLLs <1 (ig/dL remain limited.
In Anderson et al. (2012). rats were exposed starting prior to conception and then continuing
through lactation (GD -10 to PND 21) to a range of doses and assessed for learning after the end of
exposure. Only the lowest Pb concentration (250 ppm in drinking water) yielded BLLs relevant to this
ISA, with final levels of 19 (ig/dL in males and 18 (ig/dL in females. During the training phase, Pb
exposure had no effect on escape latency in either sex; however, Pb-exposed females displayed
significantly decreased path efficiency compared with untreated controls. While no effect on spatial
memory was observed during the probe trials, exposed females once again exhibited lower path efficiency
scores compared with untreated controls, suggesting that, in females particularly, Pb may influence
pathfinding processes. This study also determined that Pb partially blunted the positive effects of an
enriched environment on spatial learning, which may be relevant when considering how environmental
factors (e.g., SES) may interact with Pb exposure in humans.
In Zhao et al. (2018). rats were developmentally and lactationally exposed to multiple doses of Pb
from GD -14 to PND 10, resulting in final BLLs of 1 (ig/dL for 0.005% Pb and 1.5 (ig/dL for 0.01% Pb
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on PND 30. The highest dose group (0.02% Pb in drinking water) yielded BLLs higher than relevant for
this ISA. At the two relevant doses, Pb exposure led to significant impairments in both learning and
memory. Additional recent studies provided evidence that developmental exposure to Pb resulted in
learning and memory deficits that persisted later into adolescence and adulthood (Xiao et al.. 2020; Li et
al.. 2016a; Zhang et al.. 2014; Rahman et al.. 2012b; Zhang et al.. 2012) In one discrepant study, Wang et
al. (2021b) exposed Sprague Dawley rats to Pb in drinking water (0.05-0.2% Pb) from 4 weeks prior to
conception to PND 21. Only the lowest exposure concentration (0.05%) resulted in a mean BLL relevant
to this ISA (21.1 (ig/dL at PND 21). Learning and memory were also assessed via the Morris Water Maze
on PND 21; the authors reported no significant effects of Pb during acquisition or testing in the 0.05%
exposure group, though some effects on memory were seen at higher concentrations.
One recent study investigated the effects of Pb exposure on the cognitive function of adolescent
rodents. Liu et al. (2022c) exposed 4-week-old Sprague Dawley rats to 0.2% Pb for 12 weeks, which
yielded a mean BLL of 17.3 (ig/dL. The authors reported no effect of Pb on learning during the
acquisition phase; however, during the probe trial, Pb-exposed animals exhibited significantly fewer
crossings relative to untreated controls, suggestive of memory impairment. These recent studies provide
broadly consistent evidence that Pb produces learning and memory impairments, with developmental
periods potentially representing a more sensitive window for exposure.
Learning and Memory - Novel Objection Recognition
Another commonly applied measure of long-term memory in animal models is the novel object
recognition task. Following habituation to an empty arena, animals are placed in the arena with two
identical mundane objects and allowed to explore freely. During the testing phase (-24 hours after
training), animals are returned to the arena with one object from the first day and one novel object and
allowed to explore. The time spent examining each object is recorded. Because rodents tend to explore
unfamiliar objects, these durations can be used to calculate a recognition index, which serves as a measure
of memory. Decreased recognition indices (less time spent with the novel object) suggest impaired
memory. The previous ISA did not incorporate novel object recognition data, but one recent study used
the paradigm to assess long-term memory following Pb exposure relevant to the current assessment.
Tartaglione et al. (2020) observed that long-term exposure to Pb via maternal drinking water (GD
-28 to PND 23), which yielded relatively high BLLs of -26 (ig/dL, caused a significant decrease in novel
object recognition index in females, but not males, when tested at PND 60-72. This study did not report
results from the pre-test phase (i.e., habituation and familiarization), so the potential influence of activity
differences or inherent place preference cannot be determined. However, the result of this single study is
generally consistent with the pattern of memory impairment observed following developmental Pb
exposure, though evidence from the novel object recognition paradigm remains limited. Sex, exposure
timing, and behavioral history may also influence effects on long-term memory, yet the contribution of
each of these factors remains unclear.
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Learning and Memory-Y Maze
Another measure of spatial memory in rodents is the Y maze, which relies on the natural
inclination of rodents to explore new areas rather than revisit previously explored areas. After the animal
in placed in the Y-shaped maze, spontaneous alterations (i.e., entries into an arm different than the most
recently visited arm) and total arm entries are recorded. Total arm entries reflect locomotor activity, and
re-entries into the most recently visited arm from the center of the maze (decreased spontaneous
alteration) may indicate dysfunction in working spatial memory. The previous ISA reviewed evidence
from only one study that utilized the Y maze: (Niu et al.. 2009) reported that Wistar rats exposed to Pb
from lactation up to 12 weeks of age displayed learning impairments starting at 8 weeks of age. These
exposures resulted in BLLs of 17 (ig/dL, which are relevant to the current assessment.
Three recent studies utilized the Y maze to assess spatial memory following Pb exposures that
produced comparable BLLs (Table 3-4T), and the results were inconsistent. Xiao et al. (2020) reported
that female Sprague Dawley rats with long-term developmental exposure to Pb (125 ppm in drinking
water from GD -7 to PND 68) displayed a significant decrease in spontaneous alterations compared with
control females (70 versus 55%), which suggests a deficit in spatial working memory independent of
locomotor function. In contrast, Tartaglionc et al. (2020) did not observe any changes in spontaneous
alterations following a shorter exposure in male and female Wistar rats (50 mg/L in drinking water from
GD -28 to PND 23). Tartaglionc et al. (2020) did report a significant decrease in arm entries made by
exposed rats, which may indicate a Pb-induced alteration in exploratory behavior rather than an effect on
memory. No sex effects were observed in this study. Similarly, Abazvan et al. (2014) conducted a dietary
exposure to Pb from conception to adulthood (approximately 6 months), which yielded BLLs of 26 (ig/dL
in males and 35 (ig/dL in females (not relevant to this ISA). The authors reported no significant effect of
Pb on alterations in the Y maze in either sex. While these recent studies were focused on assessing
memory rather than learning in the Y maze, the effects of Pb on Y maze performance and the influence of
the developmental window remain unclear. Further investigation may be needed.
Learning and Memory - Fear Conditioning
Another measure of learning and memory, fear conditioning, is a task in which animals are
trained to associate a particular conditioned stimulus (e.g., auditory tone) with an aversive unconditioned
stimulus (e.g., mild foot shock). After repeated pairings of the conditioned and unconditioned stimuli
(acquisition), animals are exposed to the conditioned stimulus and the conditioned response (e.g.,
freezing, defined as lack of non-respiratory movement) is recorded. Decreases in freezing behavior may
indicate memory deficits, as the animal is no longer associating the tone with the aversive stimulus.
Several variations on this procedure may be employed to interrogate different brain regions and
processes, such as "trace" fear conditioning, wherein an interval occurs between the tone and the aversive
stimulus. Though fear conditioning data were not incorporated in the previous ISA, four recent studies
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examined the effects of Pb exposures that produced BLLs relevant to the current assessment on
associative memory.
To assess the influence of exposure window on Pb-induced cognitive impairment, Anderson et al.
(2016) exposed male and female Long-Evans rats to Pb (150, 375, and 750 ppm in chow) during three
separate exposure windows (perinatal [GD -10 to PND 21], early postnatal [PND 0 to PND 21], and
long-term postnatal [PND 0 to PND 55]), all of which resulted in BLLs <10 (ig/dL (summarized in Table
3-4T). The authors used a "trace" fear conditioning paradigm with memory testing at 1,2, and 10 days
after conditioning. Anderson et al. (2016) reported significant effects of Pb that differed by sex, exposure
window and dose. In females, learning impairments were observed only in the highest dose group of the
perinatally exposed animals. Some memory deficits were noted in the early postnatal exposure group but
only at lower doses. Interestingly, following long-term postnatal exposure, females only displayed
memory problems at the lowest doses of Pb. This result is not easily explained by variation in BLL (i.e.,
the lowest dose group did not have higher BLLs than the other dose groups). In males, minor learning
deficits were noted in the early postnatal and long-term exposure groups. Memory impairment was noted
in perinatally exposed males at the lowest and highest doses only. The results of this experiment suggest
that both sex and exposure window influence the effects of Pb on learning and memory, and that these
effects may not follow the traditional dose-response relationship reported using other paradigms.
In a subsequent study by the same group, Verma and Schneider (2017) compared the effects of
Pb on associative memory in two different rat strains using a design similar to the previous study
(Anderson et al.. 2016) to examine the influence of exposure window. There were no significant
differences in BLL between Sprague Dawley and Long-Evans rats (Table 3-4T). The authors reported no
Pb effects on acquisition across sex, indicating no effect on learning within the fear conditioning
paradigm. In Long-Evans females, animals exposed during the early postnatal period showed a marked
decrease in percent time freezing during the memory tests (day 1, control: 90% time freezing versus
treated: 68% time freezing). Consistent with the previous study from this group, the effect was more
pronounced after the initial acquisition trials (day 10, control: 70% time spent freezing versus treated:
29% time freezing). Conversely, in Long-Evans males, there was no effect in the postnatal exposure
group, yet significant impairments were detected in the perinatal exposure group starting on the 2nd day
after acquisition (day 2, control: 68% freezing versus treated: 48% freezing). Once again, the effect was
more pronounced later in the experiment (day 10, control: 75% freezing versus treated: 39% freezing).
Interestingly, no significant effect of Pb on learning or memory was observed in Sprague Dawley rats of
either sex with BLLs of approximately 5 (ig/dL.
Wang et al. (2016) exposed male Sprague Dawley rats to 100 ppm Pb in drinking water from
PND 24 to 56 and then assessed memory using a context-dependent fear conditioning paradigm in which
the rats were returned to the same test chambers without atone or shock 24 hours after acquisition, and
freezing was recorded. In this version of the test, environmental context serves as a cue that the animals
associate with the aversive stimulus. The authors reported a dramatic decrease in % time spent freezing in
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treated animals during the memory test 24 hours later (64% time spent freezing in controls compared with
only 8% time spent freezing in treated animals). This long-term adolescent exposure, which produced
BLLs of 13 |ig/dL. resulted in significant memory dysfunction. The divergent results between Verma and
Schneider (2017) and Wang et al. (2016) may be explained by variations in the paradigms used and the
duration, timing of the exposures, and resulting BLLs.
Abazvan et al. (2014) conducted dietary exposure to Pb from conception to adulthood
(approximately 6 months) in mutant (double transgenic) Disrupted-in-Schizophrenia-1 (DISCI; a genetic
risk factor for schizophrenia) mice, which yielded BLLs of 26 (ig/dL in males and 35 (ig/dL in females
(outside PECOS). Single transgenic mice that possessed the mouse DISC 1 (mDISCl) transgene but did
not express mDISCl served as controls. Using a contextual fear conditioning paradigm, these authors
reported no effect of Pb on fear extinction in mutant or control mice. These studies suggest developmental
Pb may adversely affect learning and memory within the fear conditioning paradigm, but these effects
may be sensitive to factors such as sex, strain or genetics, dose, and timing of exposure.
Learning and Memory - Avoidance
Another measure of learning and memory in animal models is the avoidance paradigm, which is a
fear-aggravated test that relies on animals learning to avoid environments where they experienced an
aversive stimulus. In the passive, "step-through" variation of the test, animals are placed in an arena with
at least two compartments, separated by gates that allow passage between compartments. During training,
animals will receive an aversive stimulus (foot shock) in the darkened chamber. The animals are later
placed in an illuminated chamber and the time that elapses before the animals enter the dark chamber is
recorded (entry latency). Shorter entry latencies are associated with impaired memory. The previous ISA
did not incorporate any passive avoidance studies. Four recent studies examined passive avoidance
behavior following Pb exposure.
Barkur and Bairv (2015b) compared the effects of Pb (0.2% in maternal drinking water) on
associative learning in male Wistar rats across several different developmental exposure periods and
durations, all but the longest of which produced BLLs <30 (ig/dL. All exposed animals, except for the
pregestation group (GD -30 to GD 1), displayed decreased entry latencies relative to controls, indicative
of impaired memory. These effects persisted out to 48 hours after the initial exploration trial. Similarly,
Barkur et al. (2011) observed that male Wistar rats exposed via maternal drinking water (0.2%) from GD
0 to PND 21 had significantly shorter entry latencies when assessed at PND 25 and again at PND 120. It
should be noted that BLLs were >30 (ig/dL when measured on PND 25 but the levels decreased to -0.5
(ig/dL by PND 120. Following long-term exposure to Pb via drinking water (50 ppm, GD 0 to PND 45),
Biioor et al. (2012) reported that male and female offspring displayed significantly shorter entry latencies
than their untreated counterparts.
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One study utilized a "step-down" version of the test, wherein animals are placed on a platform
above a grid that delivers a mild electric shock. Over the course of training, animals should learn to
associate the grid with the aversive stimulus and avoid stepping down off the platform. During testing,
both latency (time elapsed before stepping down) and errors (number of times the animal stepped down
onto the grid) are recorded to assess memory. Following long-term developmental exposure to Pb (0.4%
in maternal drinking water from GD 0 to PND 21, BLLs of -14 (.ig/dL). Kunming mice displayed
significantly decreased step-down latency and increased errors relative to controls (Zhang et al.. 2014).
suggestive of both learning and memory impairment. These studies suggest that exposure to Pb during
development results in negative effects on associative learning and memory that may persist into
adulthood and that these effects are influenced by the developmental window during which exposure
occurs.
Learning and Memory with Stress
The paradigm of combined Pb and stress exposure experienced by a laboratory animal has been
examined by the Cory-Slechta laboratory with a focus on the common pathway of an altered
hypothalamic pituitary adrenal (HPA) axis and brain neurotransmitter levels. Effects on learning varied,
depending on the timing of stress, Pb exposure concentration, and sex of the animal. Pb-stress interactions
were found with dietary Pb exposures that resulted in BLLs relevant to this ISA. The evidence
additionally indicated that associations of Pb exposure and stress with learning deficits (multiple
schedules of repeated learning and performance in females) may be related to aberrations in
corticosterone and dopamine. Several recent studies with Pb exposures relevant to the current ISA
included stress components in their experimental designs, providing further evidence that supports an
interaction between stress experience and the effects of Pb on cognitive function.
The Cory-Slechta laboratory expanded upon their previous research by investigating the
interaction between Pb and prenatal stress in males, with additional comparisons between maternal-only
and lifelong exposure (Cory-Slechta et al.. 2012). In contrast to previous reports on females, prenatal
stress with Pb exposure was reported to enhance learning accuracy with a repeated learning and
performance schedule. The authors postulated that this effect may be due to increases in the response rate,
which have been observed in both Pb and stress independently. Thus, this seemingly positive result may
reflect an increase in response rate or impulsivity.
In (Anderson et al.. 2012). researchers examined the influence of differential rearing conditions
(enriched or barren) on Morris water maze performance in Sprague Dawley rats exposed to Pb. The
authors reported a significant positive effect of the environment on learning (decreased escape latencies)
in males regardless of Pb exposure, though this effect was only present during the first two acquisition
trials. The same trend was observed in female rats, though Pb was shown to dull the advantage provided
by enrichment, with complete negation of the advantage observed at the high dose in females. While no
consistent effect was noted on time spent in the target quadrant, Pb significantly impaired the path
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efficiency relative to controls in both sexes, and this effect was ameliorated by an enriched environmental
status. While these studies on the influence of Pb and stress on learning in rodents produced variable
results, they provide evidence supporting the interaction between Pb and stress and suggest that these
effects are further influenced by sex, age, and timing of exposure.
Summary
Several recent studies in laboratory animal models with exposures resulting in mean BLLs <30
(ig/dL add to the substantial body of evidence indicating that Pb exposures can impair learning and
memory. Compared with studies in the 2013 Pb ISA, more recent studies demonstrate that the
relationship between Pb exposures and learning and memory impairments is present at lower BLLs.
Recent studies provided evidence that early-life exposures were associated with learning and memory
impairments later in adulthood, indicating that development is a critical window for the effects of Pb on
cognitive function. Additionally, new evidence suggests that longer durations of Pb exposure (especially
those encompassing developmental windows) produced greater learning and memory impairments. The
few studies reporting weak or null effects were not stronger with respect to the design or methodology
and did not weaken the much larger body of supporting evidence.
3.5.1.4 Executive Function in Children
The executive function domain of cognitive function is related to intelligence. Indices of
executive function are generally comparable to endpoints examined in experimental animal studies.
3.5.1.4.1 Epidemiologic Studies
Epidemiologic evidence presented in the 2006 Pb AQCD and the 2013 Pb ISA indicated a
consistent pattern of associations between higher childhood blood Pb (i.e., blood Pb metrics including
early childhood, lifetime average and concurrent) or tooth Pb levels reflecting pre- or early postnatal Pb
exposure, and poorer performance on tests of executive function in children and young adults (see Table
4-8 (U.S. EPA. 2013)). Associations were found with indices of executive function such as strategic
planning, organized search, flexibility of thought and action to a change in situation, and control of
impulses assessed by various tests including the Intra-Extra Dimensional Set Shift, Wisconsin Card
Sorting Test, and Stroop Color-Word test (SCWT). The strongest evidence was provided by prospective
analyses. These analyses included several birth cohorts in Boston and Rochester and examined BLLs that
preceded the outcome assessment, with adjustment for several potential confounding factors (Canfield et
al., 2004; Canfield et al., 2003b; Bellinger et al., 1994a; Stiles and Bellinger, 1993). Moderate to high
follow-up participation that was not biased to those with higher BLLs and lower cognitive function was
an additional strength of the studies. A small number of cross-sectional studies also found concurrent
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blood Pb-associated decrements in executive function, including an analysis of the Rochester cohort
(Froehlich et al.. 2007). and some studies were limited due to their lack of consideration of potential
confounding (Nelson and Espy. 2009; Vega-Dienstmaier et al.. 2006). A cross-sectional analysis by Cho
et al. (2010) with a concurrent child mean BLL of 1.9 (ig/dL did not find an association with performance
on SCWT.
A small number of recent studies expanded the evidence base pertaining to the effect of Pb on
executive functions in children (study details can be found in Section 3.7, Table 3-4E). Most of these
recent studies assessed executive function using parent-teacher ratings on the Behavior Rating Inventory
of Executive Function (BRIEF) (Gioia et al.. 2002). This instrument comprises three scales including a
Behavioral Regulation Index, which has several components (i.e., emotional control, shift, and inhibit).
Other scales of the BRIEF are the metacognition index and the global executive composite. Higher
BRIEF scores indicate executive function-related behavioral dysfunction.
Fruh et al. (2019) studied mother-child pairs participating in Project Viva, a longitudinal birth
cohort in eastern Massachusetts. Maternal blood Pb concentration in erythrocytes was measured during
the second trimester of pregnancy and parents rated their child's behavior using the BRIEF in mid-
childhood (median 7.7 years). The associations (i.e., |3 coefficients) with the parent and teacher-rated
BRIEF Behavioral Regulation Index were imprecise, i.e., 1.15 (95% CI: -0.22, 2.52) and 0.77 (95% CI:
-0.57, 2.10) per 1 (ig/dL increase in maternal erythrocyte Pb, respectively. In another analysis of these
data, Fruh et al. (2021) aimed to determine the association of joint exposure to Pb, Mn, Se, and MeHg
with scores on the BRIEF and Strengths and Difficulties Questionnaire (SDQ) using BKMR and quantile
g-computation. Individual beta coefficients for each metal from multiple regression models generally
agreed with the original results. Specifically, maternal Pb concentration in erythrocytes (2nd trimester)
was associated with worse parental ratings on the BRIEF global executive composite (|3 = 1.11 [95% CI:
(-0.12, 2.34] per unit increase in maternal erythrocyte Pb). Notably, the mixture was also associated with
poorer parent ratings on the BRIEF in BKMR models.
Sex-specific findings were observed in a study by Merced-Nieves et al. (2022). The researchers
examined the association of prenatal BLL with behavioral tasks on the operant test battery (OTB) (i.e.,
Conditioned Position Responding [CPR], Temporal-Response Differentiation [TRD], Delayed Matching-
to-Sample [DMTS], and Incremental Repeated Acquisition), which assess executive functions, at age 6-7
years. Maternal blood Pb in late pregnancy was not associated with greater response latencies in the CPR
(|3 = 0.00 [95% CI: -0.08, 0.08]) and DMTS (|3 = 0.08 [95% CI: -0.04, 0.20]) tasks, although a small
increase average latency to initiate a response in the TRD task (|3 = 0.14 [95% CI: -0.00, 0.29]). The
association of blood Pb concentrations with two operant tasks were modified by child sex, indicating Pb-
associated changes in the CPR task were more pronounced in girls, and Pb-associated changes in the TRD
task were more pronounced in boys. The mean BLLs during the first trimester, second trimester, and
delivery, in umbilical cord blood, and postnatal were 3.7, 3.9, 4.3, 3.4, and 2.4 (ig/dL, respectively.
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Ruebner et al. (2019) evaluated the association between BLLs and executive function among
children with CKD. In addition to adjusting for important potential confounders including SES and
maternal education, the author adjusted for clinical variables in their models (i.e., CKD stage, duration,
glomerular versus non-glomerular diagnosis, hypertension, proteinuria, and anemia). Executive
functioning was assessed with the Delis-Kaplan Executive Function System Tower Subset (subjects >6
years) and rated by parents using BRIEF for Preschool Children (BRIEF-P; 2-5 years) and the standard
BRIEF (6-18 years) or self-reported by adults (18 years and older) using BRIEF for Adults (BRIEF-A).
Associations between BLL and behavioral symptoms on BRIEF did not persist in models that controlled
for potential confounders including race, poverty, maternal education, and clinical factors related to CKD
(quantitative results not reported). The median BLL in this study was 1.2 (ig/dL.
Summary
Strong evidence of associations between Pb exposure and indices of executive function was
described in the 2013 Pb ISA. Studies included prospective analyses of several birth cohorts with
moderate to high follow-up rates in Boston and Rochester that examined the BLLs that preceded the
outcome assessment and adjusted for several potential confounding factors (Canfield et al.. 2004;
Canfield et al.. 2003b; Bellinger et al.. 1994a; Stiles and Bellinger. 1993). Recent studies that assessed
executive functions using parent or teacher behavioral ratings on BRIEF are mixed; however, findings
from these studies do not diminish the evidence from the earlier well-conducted studies that relied on
neuropsychological testing.
3.5.1.4.2 Toxicological Studies of Executive Function
The epidemiologic evidence reviewed above indicated associations between higher childhood
BLLs and poorer performance on tests of executive functions in children and young adults. Pb was
associated with impaired strategic planning, organized search, flexibility of thought and action to a
change in situation, and control of impulses (described in Section 3.5.1.4.1). In rodents, reversal learning
is one of the main frameworks used to measure cognitive flexibility, an important component of executive
function. Reversal learning tasks assess the ability of animals to actively suppress reward-related response
and disengage from ongoing behavior when the conditions governing the response are altered. For
example, if rodents are trained to press the leftmost of two levers to receive a reward, then the
experimenters cease to reward presses of the left lever and begin rewarding right-lever presses, the
animals must learn that the conditions have changed and adjust their behavior accordingly. Perseverance
(i.e., continuing to press the left lever after the rules have changed) represents impaired cognitive
flexibility and execution dysfunction. This basic paradigm can be expanded to parse specific components
of executive function.
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Two recent studies from the same group assessed executive function using an attention-set
shifting task paradigm following relevant Pb exposures during development. Neuwirth et al. (2019c)
compared performance in Long-Evans rats exposed to Pb during different windows of development.
Briefly, rats were trained to dig for treats by relying on environmental cues that indicated which of two
bowls contained a buried treat. Trained rats were run through a series of discrimination trials including
interdimensional and extradimensional shifts to test cognitive flexibility. An interdimensional shift occurs
when the relevant cue changes but remains within the same dimension (e.g., the baited bowl is still
indicated by a scent, but the correct scent has changed from lavender to peppermint). More complicated
extradimensional shifts require animals to recognize that the relevant cue has changed dimensions (e.g.,
the baited bowl is no longer indicated by scent but by the texture of the media).
Male rats exposed to Pb via lactation (150 ppm in maternal chow) during the early postnatal
period (PND 0 to 22), which yielded BLLs of ~6 (ig/dL, displayed substantial learning deficits in the form
of increased Trials-to-criterion for the olfactory (the relevant dimension) discrimination component.
Indeed, males exposed during the early postnatal period were unable to complete discrimination training
and progress to the next task in the same manner as the control males, indicative of a substantial learning
impairment as the result of Pb exposure. No effect was observed in female rats following exposure in the
postnatal period, despite similar BLLs of ~5 (ig/dL. Even though male rats perinatally exposed (GD -14
to PND 22, BLLs of ~6 (ig/dL) successfully completed discrimination training, they struggled during
testing and displayed significant increases in Trials-to-criterion across simple and complex discrimination
tasks yet solved extradimensional shift tasks in a fashion comparable to control males. In contrast to
males, perinatally exposed females performed poorly in extradimensional shift tasks and showed
improved performance relative to controls in discrimination training and reversal (Neuwirth et al.. 2019c).
In a subsequent study by the same group (Neuwirth et al.. 2019b). the authors reported that
perinatal exposure (GD 0 to PND 22), which resulted in BLLs -10 (ig/dL on PND 22, also produced sex-
specific learning deficits in discrimination training and impaired reversal learning. These studies provide
evidence that exposure to Pb during different developmental windows may produce differential patterns
of executive dysfunction and that these changes may be sex-specific. While the previous ISA did not
include any toxicological evidence that explicitly addressed executive function, the findings of (Neuwirth
et al.. 2019c) and (Neuwirth et al.. 2019b) are consistent with the evidence reviewed in the previous ISA
that indicated Pb exposure contributed to cognitive dysfunction and that the effects of Pb on cognition
were often sex-specific.
Summary
The previous ISA did not incorporate any evidence of the relationship between Pb exposure and
executive function in animal models. Two recent studies from the same group provided evidence that Pb
exposure broadly impairs measures of executive function in a reversal learning paradigm. These effects
were sex-specific, with greater effects reported in males. While these reports are consistent with one
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another, evidence for the association between Pb exposure and impaired executive function in animal
models with BLLs <30 (ig/dL remains limited.
3.5.1.5 Academic Performance and Achievement in Children
Poorer academic performance and achievement is linked with lower FSIQ and may have
important implications for success later in life (U.S. EPA, 2013). The 2006 Pb AQCD and the 2013 Pb
ISA described associations of higher blood and tooth Pb levels, which reflected Pb exposure at various
time periods and lifestages, in children aged 5-18 years with poorer performance on tests of math,
reading, and spelling skills, lower probability of high school completion, lower class rank, and lower
teacher ratings of academic functioning. Notably, associations were reported in prospective studies
examining performance on academic achievement tests (Chandramouli et al., 2009; Min et al., 2009;
Miranda et al., 2009) and an additional analysis of adolescents participating in NHANES (Lanphcar et al..
2000). Several prospective studies (Min et al., 2009; Miranda et al., 2009) and cross-sectional studies
(Krieg et al., 2010; Chiodo et al„ 2007; Surkan et al„ 2007; Lanphear et al„ 2000) were conducted in
populations with population or group mean BLLs <5 (ig/dL. In addition, prospective studies in Boston
and New Zealand found associations of tooth Pb levels collected at an earlier age (e.g., ages 6-8 years)
and generally reflecting pre- or early postnatal Pb exposure, with school performance ascertained at age
18 from school records (Fergusson et al., 1997; Needleman et al„ 1990), suggesting the effect of early
exposure on Pb may be persistent. The strengths of the Fergusson et al. (1997) analysis included a low
probability of selection bias, coherence with results indicating associations between higher tooth Pb levels
and lower teacher ratings of math, reading, and writing abilities at ages 12-13 years (Fergusson et al.,
1993), and consideration of important covariates including SES, parental education, and HOME score.
The Needleman et al. (1990) study was relatively small with no adjustment for parental caregiving
quality.
Recent prospective studies of groups or populations with mean BLLs <5 (ig/dL add to the
evidence supporting an effect of Pb exposure on academic achievement and performance (study details
can be found in Section 3.7, Table 3-5E). Prospective studies have been conducted in Detroit, Chicago,
North Carolina, and a 57-county region in New York State (all counties outside New York City).
Zhang et al. (2013) studied children enrolled in Detroit public schools to determine the
association between childhood BLL measured before age 6 and performance on standardized tests for
math, reading, and science in grades 3, 5, and 8. Compared with students with lower BLLs (defined as
levels <1 (ig/dL), students with higher BLLs (defined as levels between 1 and 5 (ig/dL) had increased risk
of scores that were classified as less than proficient (OR = 1.42 [95% CI: 1.24, 1.63] for math, OR = 1.33
[95% CI: 1.10, 1.62] for science, and OR = 1.45 [95% CI: 1.27, 1.67] for reading. Logistic regression
models were adjusted for covariates including SES (i.e., free and reduced school lunch participation) and
maternal education.
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Evens et al. (2015) conducted a study in children enrolled in the Chicago public school system.
The association between childhood BLLs measured before 72 months and failure on standardized tests for
math and English in grade 3 was examined. This study found that a 1 (ig/dL increase childhood BLL was
associated with an increased risk of failure on the reading and math tests (RR = 1.06 [95% CI: 1.05, 1.07]
and RR = 1.06 [95% CI: 1.05, 1.07], respectively) after adjustment for covariates including child
characteristics, preterm birth, maternal education, and SES (i.e., participation in the free and reduced
lunch program). The associations of BLL with reading failure in white, Black and Hispanic children were
1.14 (95%CI: 1.08, 1.20), 1.05 (95% CI: 1.04, 1.06) and 1.08 (95%CI: 1.05, 1.11), respectively. The
associations of BLL with math failure were in white, Black and Hispanic children were 1.11 (95% CI:
1.05, 1.18), 1.05 (95% CI: 1.04, 1.06) and 1.09 (95% CI: 1.06, 1.12), respectively. The mean BLL was
4.81 (ig/dL in this study. Blackowicz et al. (2016) extended this analysis through their examination of the
association of BLL and failure on standardized tests for math or reading among Hispanic children
enrolled in the Chicago school system. An association between 1 (ig/dL change in BLL and failures in
reading (RR= 1.07 [95% CI: 1.05, 1.10]) and math (RR= 1.09 [95% CI: 1.06, 1.12]) were observed. The
mean BLL was 4.16 (ig/dL in Hispanic children in this study.
In a statewide study of North Carolina school children (Shadbegian et al.. 2019). children with
higher BLLs had, on average, lower scores in both math and reading (averaged over grades 3 and 8) than
children with lower BLLs. Compared with children with BLL <1 (ig/dL, the authors reported a decrease
in the test-score percentile of 0.95 (0.66, 1.24) for math and 1.41 (1.12, 1.70) for reading in children with
a BLL of 5 (ig/dL. Shadbegian et al. (2019) included interaction terms between BLL and the grade of
testing that further indicated that the deficit in the test score persisted from grade 3 to grade 8.
SkerfVing et al. (2015) studied the association of childhood BLL (age 7-12 years) with school
performance in the ninth grade at age 16 among Swedish school children. School performance was based
on a 1-5 point passing grade scale or a 4-level merit system in which 0, 10, 15, or 20 points were
assigned for each increasing level of performance. This study found a 0.11-point decrease (95% CI:
-0.18, -0.05) per 1 (ig/dL increase in BLL for school performance using the grading scale and a -10.90
(95% CI: -15.49, -6.31) point decrease using the merit scale, among school children with BLLs <5
(ig/dL. The models were adjusted for covariates including parent's income, education, and father's IQ
score on the military conscription exam. The association with IQ (|3 = -0.20 [95% CI: -0.39, -0.02])
among those evaluated for military conscription at age 18 was also examined (see Section 3.6.1).
3.5.1.5.1 Summary
Associations of higher blood and tooth Pb levels, which reflect various time periods including
earlier childhood lifestages, in children aged 5-18 years with poorer performance on tests of math,
reading, and spelling skills, lower probability of high school completion and lower-class rank, and
lower teacher ratings of academic functioning were observed in previous assessments (U.S. EPA, 2013).
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Recent studies in populations of children (age 6-16 years) with BLLs <5 (ig/dL support and extend these
observations of poorer academic performance in association with increasing Pb exposure.
3.5.1.6 Relevant Issues for Interpreting the Evidence Base
3.5.1.6.1 Concentration-Response Function
With each previous assessment (U.S. EPA, 2013, 2006), the epidemiologic and toxicological
study findings have shown that progressively lower BLLs or Pb exposures are associated with cognitive
deficits in children. The 2006 AQCD found that cognitive effects in children were associated with BLLs
of 10 (ig/dL and lower, while the evidence assessed in the 2013 Pb ISA found that an association between
BLLs and cognitive effects in children was substantiated to occur in populations of young children with
mean BLLs between 2 and 8 (ig/dL. The conclusions of the 2013 Pb ISA were based on studies that
examined early childhood BLLs (i.e., age <3 years), considered peak BLLs in their analysis (i.e., peak
BLL <10 (ig/dL), or examined concurrent BLLs in young children (i.e., age 4 years). The lower bound of
this mean BLL range was derived from Miranda et al. (2009), who examined the association between
early childhood BLL and academic performance among school-aged (grade 4) children. A recent study of
Canadian preschool children from generally middle- to upper-middle SES families with low Pb exposure
(Desrochers-Couture et al., 2018) did not find an association between concurrent Pb exposure and
performance on the WPPSI at age 3-4 years. Although some individual recent studies found associations
of Pb exposure with cognitive effects in children with mean BLLs <2 (ig/dL (e.g., (Martin et al„ 2021;
Dantzer et al.. 2020; Hong et al„ 2015)). the studies generally involved somewhat older children with
lengthier exposure histories, or employed modeling strategies designed to answer relatively narrow
research questions (e.g., the effect of joint exposure to Pb and other metals or the effect of concurrent Pb
exposure independent from prenatal exposure). Consequently, the studies did not provide evidence that
would change the conclusion of the 2013 Pb ISA that cognitive effects in children are best substantiated
in young children with mean BLLs between 2 and 8 (ig/dL. Studies that might extend the evidence related
to exposure-response relationships (i.e., recent studies that reflect the lower early childhood Pb exposures,
which are now more common in the U.S.]) are limited. Overall, the recently available studies were not
designed, and may not have the sensitivity (Cooper et al., 2016), to detect the effect or hazard at very low
BLLs; however, recent studies generally corroborated the epidemiologic observations of associations
between Pb exposure and IQ in children with relatively low blood Pb concentrations (<5 (.ig/dL).
Consistent with findings from the 2013 Pb ISA, studies do not provide evidence of a threshold for the
effects across the range of BLLs examined. The finding of higher mean IQ with decreasing blood Pb
concentration observed across epidemiologic studies, however, indicates that the absolute magnitude of
the effect of Pb exposure on cognitive function is smaller with decreasing BLL.
Despite limitations, several recent studies describe the cognitive effects over the range of Pb
exposure examined. Shadbcgian et al. (2019) extended the analysis conducted by Miranda et al. (2009)
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also using data from the statewide study of North Carolina school children while focusing on lower BLLs
(<10 (ig/dL) and characterized the persistence of Pb effects across grades. Among children with BLLs of
5 (ig/dL or lower, a decrease in the test score percentile of 0.95 was found for math and a decrease of 1.41
was found for reading when comparing children with a BLL of 5 to those with a BLL <1 (ig/dL. Figure
3-6 depicts the association of blood Pb levels with math and reading test score performance (average
percentile decrement), with 95% CIs, among children across all grades.
Dotted lines denote 95% confidence intervals.
Source: CShadbeqian et al.. 20191.
Figure 3-6 Association of blood Pb level with reading and math scores
among North Carolina school children (average across all
grades). Left panel displays impact of blood Pb level on math test
score. Right panel displays impact of blood Pb level on reading
test score.
Compelling evidence for a larger decrement in cognitive function per unit increase in blood Pb
among children with lower mean blood Pb concentrations, compared with children with higher mean
blood Pb concentrations, was presented in previous assessments (U.S. EPA, 2013, 2006). Individual
studies as well as an international pooled analysis of seven prospective cohort studies by Lanphear et al.
(2019, 2005) that examined prenatal or early childhood BLLs or considered peak BLLs in school-aged
children or concurrent BLLs in young children <3 years old showed greater decrements in cognitive
function per unit increase in BLL among children in lower strata of blood Pb levels compared with
children in higher strata of blood Pb level (Figure 4-15, and Table 4-16 of U.S. EPA (2013) corroborated
the finding of a nonlinear C-R function over the range of the BLLs evaluated (5th to 95th percentile BLL,
2.5 to 33.2 (ig/dL), i.e., a larger incremental effect of Pb exposure on IQ at lower blood Pb concentrations
as indicated by a log-linear C-R function (Crump et al„ 2013). Notably the reanalysis by Crump et al.
(2013) extended the findings of the original study by employing a different modeling strategy.
Specifically, several covariates were defined in a site-specific manner (i.e., HOME score, maternal
education, maternal IQ, ethnicity, maternal alcohol consumption, and maternal smoking), which enabled
finer scale control for potential confounding factors (e.g., amount of alcohol consumed as opposed to a
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binary variable for alcohol consumption). In addition, Crump et al. (2013) used more of the available
blood Pb measurements by computing weighted averages for lifetime and early childhood BLLs. Further,
the authors used formal methods (i.e., nested F-tests and splines) to decide between alternative
specifications of BLL (linear versus log-linear) and chose to add 1 to the BLL prior to log transformation
to ensure it would equal zero when BLL was zero. The key findings of Lanphear et al. (2005), Lanphear
et al. (2019), and Crump et al. (2013) contribute to the strong evidence regarding the effect of low-level
Pb exposure on cognitive function and the supralinear concentration relationship between Pb exposure
and FSIQ. The beta coefficients from the log-linear models, which indicate larger incremental effects of
Pb at lower blood Pb concentration, for (Lanphear et al., 2019, 2005) and Crump et al. (2013) were
comparable (i.e., |3 = -2.65 [95% CI: -3:69, -1:61] per unit of natural log transformed BLL and |3 = -3.32
[95% CI: -4.55, -2.08] per unit of natural log transformed BLL + 1, respectively).
Attenuation of C-R relationships at higher exposure or dose levels has been reported in the
occupational literature. Reasons proposed to explain the attenuation include greater exposure
measurement error and saturation of biological mechanisms at higher levels as well depletion of the pool
of susceptible individuals at higher exposure levels (Stayner et al„ 2003). Possible explanations specific
to nonlinear relationships observed in studies of Pb exposure in children include a lower incremental
effect of Pb due to covarying risk factors such as low SES, poor caregiving environment, and higher
exposure to other environmental factors (Schwartz, 1994a), differential activity of mechanisms at
different exposure levels, and confounding by omitted or mis-specified variables (U.S. EPA, 2013).
Review of the evidence did not reveal a consistent set of covarying risk factors to explain the differences
in blood Pb IQ C-R relationship across high and low Pb exposure groups observed in epidemiologic
studies. Recent studies in populations with mean concentrations of 2 (ig/dL or lower indicated that some
of the observed heterogeneity at lower BLLs may be explained by the underlying distribution in at-risk
factors, including other metals. In addition, some recent studies with similarly low blood Pb
concentrations reported effect modification by sex. These studies are discussed in more detail in Section
3.5.1.6.2.
A limited number of recent studies examine the shape of the C-R function for the relationship
between Pb exposure and cognitive effects in children. Lucchini et al. (2012) conducted a cross-sectional
analysis of children between the ages of 11 and 14 to examine the relationship between concurrent BLL
and FSIQ. The mean BLL was 1.71 (ig/dL in this study. The relationship between BLL and IQ using a
restricted cubic spline fit is plotted in Figure 3-7. As shown in the plot, the decrement in IQ is not
constant over the range in BLLs (0.44-10.2 (.ig/dL). The study was conducted in an area where ferroalloy
plants had operated and the extent to which the children in the study were exposed to higher Pb levels
during early childhood was not clear from this cross-sectional analysis.
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2
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Pb (U0ML)
6
10
IQ = intelligence quotient.
Source: Lucchini et al. (2012).
Figure 3-7 Relationship between concurrent blood Pb level and intelligence
quotient among Italian adolescents using a cubic spline fit.
Lucchini et al. (2012 also plotted relationship between the log-transformed BLL (ordinary least
squares fit) and FSIQ (Figure 3-8). A decrement in FSIQ score was observed in association with In
concurrent BLL after adjustment for covariates including SES and maternal education (-2.24 [95% CI:
-4.10, -0.37)]. Consistent with evidence reviewed in the 2013 Pb ISA, the log transformation of BLL
implies a larger incremental decrement m IQ at lower BLLs. Lucchini et al. (2012) calculated the
benchmark dose (BMD) for blood Pb, which is the dose that results in a specific IQ loss (i.e., a loss of one
IQ point), and its lower 95% confidence limit (BMDL) using this C-R function. The BMDL calculated
from these data is 0.11 |0g/dL. As noted previously the older children in this study may have had higher
past Pb exposure that was not reflected in their concurrent BLL.
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IQ = intelligence quotient.
Source; Lucchini et al. (20121.
Figure 3-8 Relationship between log-transformed blood Pb level and
intelligence quotient using an ordinary least squares fit.
Experimental animal studies support the findings in epidemiologic studies indicating the effect of
Pb exposure on cognition at low exposures. Clear support from animal toxicological studies that
demonstrated decrements in learning, memory, and executive function with dietary exposures resulting in
relevant BLLs was assessed in the 2013 Pb ISA. Recent experimental animal studies further support
impairments in cognitive function at BLLs <20 |ig/dL. It is well-documented (and reviewed in the
previous ISA) that Pb exposures resulting in BLLs >20 (ig/dL consistently produced deficits in cognitive
function. Recent evidence (reviewed in the current ISA; see Section 3.5.1.3.2) suggests that BLLs
resulting from lower-level exposures (5—10 (ig/dL) also lead to cognitive function deficits in animal
models. For example, Zhou et al. (2020a); Zhao et al (2018); Xiao et al. (2014); Bethana and Malier
(2012); Corv-Slechta et al. (2012) all reported significant cognitive deficits following exposures yielding
BLLs < 10 ng/dL. Betharia and Maher (2012) reported the lowest BLLs for animals tested in the Morris
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water maze (0.98 (ig/dL at PND 21 and later 0.03 (ig/dL at PND 56) which produced mild effects on
memory in females at the later time point only.
The reader should be aware that BLLs reported in many of these studies were measured later
during experimentation and do not necessarily reflect peak Pb burden or Pb burden during the most
sensitive window of brain development for young animals and children; BLLs in toxicological studies
should be interpreted in the context of both exposure time course and blood collection. Discussion of
BLLs and cognitive function is further complicated by the exposure window. The evidence generally
supports the notion that Pb exposures during brain development led to changes in cognition that persist
after cessation of exposure and BLLs have decreased. For example, Barkur and Bairy (2015b) compared
multiple different windows of exposure and reported the greatest decreases in learning and memory
following gestational and lactational windows, consistent with the altricial nature of rodent brain
development.
While nonlinear C-R relationships including U- or inverted U-shaped curves for various
endpoints, including those related to cognitive impairment, were demonstrated in the toxicological
literature discussed in the previous ISA, these toxicological findings are distinct from epidemiologic
findings of supralinear relationships in that some U- or inverted U-shaped relationships do not indicate
Pb-induced impairments at higher exposure concentrations (U.S. EPA, 2013). Recent animal studies do
not provide evidence for an inverted dose relationship, rather increased Pb doses (and resulting BLLs)
generally resulted in greater cognitive impairment. This may be related to the refined PECOS used in the
current ISA, which did not incorporate studies reporting BLLs higher than 30 (ig/dL, thus narrowing the
range of doses integrated. Thus, recent evidence generally supports dose-dependent effects of Pb on
cognitive function at relevant BLLs.
In summary, recent studies support and extend the evidence pertaining to the effect of Pb
exposure on cognitive function in children at low BLLs. These effects are best substantiated to occur in
study populations with mean BLLs between 2 and 8 (ig/dL. Association between Pb exposure in
populations of children below 2 (ig/dL are reported, extending the evidence described in the 2013 Pb ISA;
however, heterogeneity at lower exposure levels (i.e., not all studies report positive associations) has been
observed. Recent experimental studies of rodents continue to support impairments in cognitive function at
BLLs <30 (ig/dL. Compelling evidence for a larger decrement in cognitive function per unit increase in
blood Pb among children with lower mean blood Pb concentrations, compared with children with higher
mean blood Pb concentrations, across a broad range of BLLs (e.g., 5th percentile of 2.5 (ig/dL up to 95th
percentile of 33 (ig/dL) was supported by a reanalysis of a pooled international dataset Crump, 2013,
3838553}. Recent studies with an adequate range of Pb exposure measured during relevant time periods
that would be required to evaluate exposure-response relationships are generally lacking. Considering the
collective body of studies, no evidence of a threshold for cognitive effects in children across the range of
BLLs examined in epidemiologic studies was reported.
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3.5.1.6.2 Confounding
The 2013 Pb ISA described multiple factors that influence cognitive function and behavior in
children including parental IQ and education, SES of the family, quality of the caregiving environment
(i.e., HOME score), and other environmental exposures (U.S. EPA, 2013; Wasserman and Factor-Litvak,
2001). These other risk factors often are correlated with blood, tooth, and bone Pb levels, and thus, are
considered as potential confounding factors in epidemiologic analyses. The collective epidemiologic
evidence consistently demonstrates associations of higher blood and tooth Pb levels with cognitive
function decrements and poorer behavior in children. These associations were observed in diverse
populations in the U.S., Mexico, Europe, Asia, and Australia. Associations have been observed across
studies that used different methods to control for confounding and adjusted for different potential
confounding factors, commonly maternal IQ and education, SES, and HOME score. Several studies have
found associations with additional adjustments for smoking exposure, birth outcomes, and nutritional
factors. Multiple recent studies adjusted for exposure to other metals or environmental chemicals e.g.,
Zhou et al. (2020b) and Liu et al. (2015); however, there remains uncertainty regarding the
appropriateness of the adjustment for other metals as confounders in some studies that did not examine
the potential for interactions (see Section 3.5.1.6.5 for evidence related to interactions between Pb and
other metals).
As noted in the 2013 Pb ISA, no single method to control for potential confounding is without
limitation, and there is potential for residual confounding by unmeasured factors. However, the
consistency of findings among different populations and study methods with consideration of several well
characterized potential confounding factors as described above increases confidence that the associations
observed between Pb biomarker levels and neurodevelopmental effects in children represented a
relationship with Pb exposure. Recent studies expanded the evidence reporting associations between Pb
exposure and nervous system effects in children after consideration of covariates including sex, maternal
stress and race or ethnicity as potential effect modifiers as opposed to confounders (see Section 3.5.1.6.5).
Biological plausibility was derived from extensive evidence provided by animal toxicological studies that
are experimental in design and thus, not vulnerable to confounding. These experimental animal studies
demonstrate the effect of Pb on cognition and behavior as well as changes in neurogenesis, synaptic
pruning, and neurotransmitter function in the hippocampus, prefrontal cortex, and nucleus accumbens of
the brain (U.S. EPA, 2013). Recent experimental animal studies support the evidence described in the
2013 Pb ISA, provide additional evidence for Pb-induced impairments in learning and memory (short and
long-term) assessed by several methods not discussed in the 2013 Pb ISA, and extend the limited
evidence related to Pb-induced impairment of executive functions. These experimental animal studies
provide strong support that the effects observed in epidemiologic studies cannot be explained by
confounding.
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3.5.1.6.3 Lifestages
Epidemiologic studies reviewed in the 2013 Pb ISA consistently showed that BLLs measured
during various lifestages and time periods (i.e., prenatal, early childhood, childhood average, and
concurrent with the outcome) were associated with cognitive function decrements in children (U.S. EPA,
2013). Epidemiologic studies consistently pointed to inverse associations between FSIQ in school-aged
children and BLLs measured at various lifestages and time periods (Table 4-14 U.S. EPA (2013)). In an
analysis of data from seven prospective studies Lanphear et al. (2019) found that increases in early
childhood (age 6-24 months on average), peak, concurrent, and lifetime average BLLs were associated
with decreases in FSIQ in children at ages 4-10 years. The investigators reported that the best predictor of
IQ decrement, as indicated by the model R2 value, was early childhood blood Pb concentration (R2 =
0.6433), although the R2 value for the concurrent metric (0.6414) was nearly identical (Lanphear et al„
2019; Crump et al., 2013). These results illustrated the challenge of distinguishing a critical time period
when exposures are highly temporally correlated. Epidemiologic studies that aimed to improve the
characterization of important lifestages and time periods of Pb exposure by examining children in whom
BLLs were not strongly correlated with exposure over time indicated FSIQ decrements in association
with higher concurrent BLLs but did not conclusively demonstrate stronger findings for early versus
concurrent BLLs (Table 4-15 of U.S. EPA (2013)). Considering the collective body of epidemiologic
evidence reviewed in the 2013 Pb ISA, there was no clear indication of a single critical lifestage or
duration of Pb exposure that is uniquely associated with the risk of neurodevelopmental effects in
children. These observations in the epidemiologic literature were supported by experimental animal
evidence. Consistent with findings from the 2013 Pb ISA, more recent studies continue to report
associations with prenatal BLLs (maternal and cord blood Pb) and postnatal BLLs measured at various
childhood lifestages despite some heterogeneity in the magnitude and direction of the associations at
BLLs <5 (ig/dL.
Maternal Pb exposure presents an exposure risk during gestation and early infancy, when
important neurodevelopmental processes are known to occur. Substantial fetal Pb exposure may occur
from mobilization of maternal skeletal Pb stores (Gulson et al., 2003; Hu and Hernandez-Avila, 2002)
and its transfer across the placenta (Section 3.2.2.4 of U.S. EPA (2013)). Among studies that examined
BLLs at multiple time periods, some found a larger decrement in MDI per unit increase in prenatal blood
Pb than concurrent blood Pb ((Hu et al„ 2006; Gomaa et al„ 2002), Table 4-14 of U.S. EPA (2013)).
Prenatal and early postnatal (age 6 months) BLLs were also associated with cognitive function in studies
that included school-aged children (ages 5-17 years) (Table 4-14 of U.S. EPA (2013)). Sanchez et al.
(2011) extended the analysis of Hu et al. (2006), which was designed to elucidate the time window during
pregnancy that the effect of Pb exposure on neurodevelopment is most pronounced among participants in
a birth cohort study in Mexico City. These authors compared methods to model exposure and found that
the MDI score at age 2 was sensitive to the choice of method. A decrease in MDI score of 2.74 (95% CI -
5.78 to 0.29) per natural log increase in BLL during the first trimester was observed, using a window-
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specific regression, while the corresponding decrease was larger [|3=—4.13 (95% CI, -7.54, -0.72) using a
multiple informant model.
As described above, however, most of these studies also found cognitive function decrements in
association with postnatal BLLs, and the results did not identify an individual critical postnatal time
period of blood Pb measurement associated with cognitive function decrements. Maternal pregnancy-cord
BLL correlations of 0.53-0.81, depending on the stage of pregnancy, were reported by Schell et al.
(2003). Depending on the magnitude of child exposure, the contribution of maternal blood Pb to child
BLLs appears to diminish rapidly over a period of a few months following birth, after which child BLLs
may be influenced mainly by postnatal Pb exposures (Section 3.4.1 of U.S. EPA (2013)).
Recent studies observed associations between Pb exposure during prenatal and childhood
lifestages (i.e., maternal (Y Ortiz et al„ 2017; Vigeh et al., 2014; Kim et al„ 2013b, c), cord (Valeri et al„
2017), and postnatal exposure (Lin et al., 2013)) and poorer performance on tests of neurodevelopment
among mothers and infants with mean BLLs <5 (ig/dL. The is some evidence indicating that there
heterogeneity in the magnitude and direction of the observed associations in recent studies may be
explained, in part, by co-exposure to other metals or maternal stress
Experimental animal studies demonstrated that prenatal or early postnatal or lifetime Pb exposure
alters brain development via changes in synaptic architecture and neuronal outgrowth, leading to
impairments in memory and learning (Sections 4.3.10.4, 4.3.10.10, and 4.3.2.3 of U.S. EPA (2013)).
Gestational or infancy Pb exposures are not necessary to induce cognitive function decrements in juvenile
animals, however. Studies of monkeys have found that Pb exposures during lifestages and time periods
extending from infancy through the juvenile or adult periods resulted in impaired cognitive function
(Rice, 1992; Rice and Gilbert, 1990a; Rice, 1990; Rice and Karpinski, 1988). These findings are
consistent with studies of individuals aged 3 to 30 years, which showed that brain development
ascertained using MRI continues throughout adolescence, indicating the potential for alterations to
neurodevelopment later in childhood (Gerber et al., 2009; Lenroot and Giedd, 2006).
Additional recent animal studies support the notion that various exposure periods (i.e.,
preconception, gestation, lactation) may represent a critical periods during which Pb exposure can cause
cognitive impairment later in life. In rodents, developmental exposure to Pb was consistently associated
with persistent cognitive effects observed both early (Tartaglionc et al., 2020; Zhao et al„ 2018; Barkur
and Bairy, 2015b; Anderson et al., 2012) and later in life (Liu et al., 2022c; Xiao et al„ 2014; Betharia
and Maher, 2012). Few studies were designed to compare exposures across multiple different
developmental windows (Barkur and Bairy, 2015b; Xiao et al., 2014). These studies reported similar
magnitudes of effects between developmental windows, suggesting that individual periods of
development may be similarly sensitive to Pb. Generally, longer exposures that spanned multiple
developmental periods (e.g., preconception through lactation) produced not only the highest BLLs but the
largest effects on cognition (Zhou et al„ 2020a; Zhu et al„ 2019b).
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Unlike other organ systems, the unidirectional nature of CNS development limits the capability of
the developing brain to compensate for cell loss, and environmentally induced cell death can result in a
permanent reduction in cell numbers (Bayer, 1989). Hence, when normal development is altered, the early
effects have the potential to persist into adult life even in the absence of concurrent exposure, magnifying
the potential public health effects. A limited number of studies examined the persistence of the effects of
Pb on cognitive function. A recent study by Shadbegian et al. (2019) indicated that poorer performance on
tests of reading and math associated with earlier childhood Pb exposure persisted from grade 3 to grade 8.
Some epidemiologic evidence reviewed in the 2013 Pb ISA indicated associations of earlier childhood
blood or tooth Pb levels in adolescents or adults with decreased cognitive function (Mazumdar et al.,
2011; Ris et al., 2004; Stiles and Bellinger, 1993). Recent studies support and extend this evidence.
Specifically, childhood Pb exposure was observed to have long-term cognitive consequences in young-
(18-19 years) (SkerfVing et al., 2015) or mid-adulthood (38 or 45 years of age) (Reuben et al., 2020;
Reuben et al„ 2017). These epidemiologic studies did not examine adult BLLs, thus the relative influence
of adult Pb exposure was not ascertained. The persistence of effects of early exposures, however, is
supported by findings of impaired learning in adult monkeys exposed to Pb only during infancy (Rice,
1992; Rice and Gilbert, 1990a; Rice, 1990). Additional recent studies in rodents provided support for the
persistence of effects of early exposures of Pb (Xiao et al., 2020; Li et al., 2016a; Xiao et al., 2014; Zhang
et al., 2014; Rahman et al., 2012b; Zhang et al„ 2012; Kuhlmann et al., 1997).
There is some evidence that the effects of early Pb exposure on cognitive function are not fixed.
Results indicated higher cognitive function in children at ages 1-8 years who had declines in BLL over
durations of 6 months to 5 years compared with children with smaller declines, no change, or increases in
BLLs in some studies (Hornung et al„ 2009; Chen et al., 2005; Liu et al., 2002; Ruff et al„ 1993;
Bellinger et al„ 1990). This evidence pertains to populations with declines from higher BLLs at baseline
(20-55 (ig/dL) or larger declines over time (i.e., 8, 14 (ig/dL) than those expected for most of the current
population of U.S. children. No recent studies that provided additional information on this topic were
identified.
To conclude, the collective body of epidemiologic evidence reviewed in the 2013 Pb ISA did not
provide strong evidence to identify an individual critical lifestage or timing of Pb exposure with regard to
neurodevelopmental effects in children (U.S. EPA, 2013). Recent studies support this conclusion.
Evidence indicates that prenatal BLLs are associated with mental development in very young children
aged <2 years. Several studies indicated that increases in postnatal (earlier childhood, lifetime average,
concurrent) BLLs were associated with larger cognitive function decrements in children aged 4-10 years
than were similarly sized increases in prenatal BLLs. These results suggest that per unit increase,
postnatal Pb exposures that are reflected in concurrent or cumulative BLLs or tooth Pb levels may have a
larger magnitude of effect on cognitive function decrements as children age (U.S. EPA, 2013). The
identification of critical lifestages and time periods of Pb exposure is complicated by the fact that BLLs in
older children, although affected by recent exposure, are also influenced by Pb stored in their bone and
maternal Pb stores. Thus, associations of neurodevelopmental effects with concurrent BLL in children
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may reflect the effects of past and recent Pb exposures. Nonetheless, the epidemiologic evidence for
associations of neurodevelopmental effects with multiple lifestages or time periods of Pb exposure,
including more recent exposures, is supported by evidence in monkeys that Pb exposures in infancy,
lifetime exposure starting from birth, or lifetime exposure starting during the juvenile period induce
impairments in cognitive function when assessed between the ages of 6 and 10 years.
3.5.1.6.4 Public Health Significance
The 2006 Pb AQCD and the 2013 Pb ISA (U.S. EPA. 2013. 2006) concluded that
neurodevelopmental effects in children were among the effects best substantiated as occurring at the
lowest BLLs. Evidence from several cohorts of children indicated that there is a supralinear C-R
relationship between blood Pb and FSIQ (i.e., larger incremental effect of blood Pb on FSIQ at lower
levels) and no threshold was identified for Pb-associated neurodevelopmental effects in the range of BLLs
examined. The evidence reviewed in the current assessment supports these conclusions and continues to
clearly indicate that neurodevelopmental effects in children are among the greatest public health concern
associated with Pb exposure.
Cognitive function in children has been assessed using a variety of tests, including FSIQ, BSID,
academic performance, and academic achievement. As noted in the 2013 Pb ISA (U.S. EPA, 2013), FSIQ
has strong psychometric properties (i.e., reliability, consistency, validity), is among the most rigorously
standardized cognitive function measures, is relatively stable in school-age children, and has been
predictive of educational achievement and life success. Variation in IQ score across different populations,
however, may be influenced by differential access to resources in those populations
(Shuttleworth-Edwards, 2016; Marks, 2010). In children aged 6 months to 3 years, the BSID is commonly
used to assess mental development; however, the BSID MDI is not an intelligence test and MDI scores
are not necessarily strongly correlated with later measurements of FSIQ in children with normal
development. Lower FSIQ is also linked to poorer academic performance and achievement, both of which
have important implications for success later in life including reduced earning potential and productivity
(Lin et al., 2016; U.S. EPA, 2013; Salkever, 1995; Schwartz, 1994b). Analyses of end-of-grade tests from
North Carolina indicated that early childhood BLL is associated with reduced performance on the tests,
the cumulative effect of Pb and low SES is more pronounced at the lower end of the test score distribution
(Miranda et al„ 2009), and the effects of Pb exposure persisted from grade 3 to grade 8
(Shadbegian et al„ 2019). Tests of academic achievement generally measure a child's understanding of a
given curriculum that is developed and implemented through the school system; thus, because exams are
typically specific to each state, data cannot be directly compared across states.
The World Health Organization (WHO) definition of "health" is "the state of complete physical,
mental, and social well-being and not merely the absence of disease or infirmity" (WHO, 1948). By this
definition, decrements in health status that are not severe enough to result in the assignment of a clinical
diagnosis might reflect a decrement in the well-being of an individual. Further, deficits in subtle indices
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of health or well-being may not be observable except in aggregate, at the population level; therefore, a
critical distinction between population and individual risk is essential for interpreting the public health
significance of study findings. This concept of population risk is relevant to the interpretation of findings
regarding IQ in the assessment of their public health significance. Specifically, Weiss (1988) discussed
the hypothetical effects of a small shift in the population distribution of IQ score. As shown in Figure 3-9,
these authors anticipate that even a small shift in the population mean IQ may be significant from a public
health perspective because such a shift, given a normal distribution, could yield a larger proportion of
individuals functioning in the low range of the IQ distribution, which is associated with increased risk of
educational, vocational, and social failure (Section 4.3.13), as well as reduce the proportion of individuals
with high IQ scores. Although the change in population mean IQ score may be small relative to the
standard error for the IQ measurement, a study that is large enough will have adequate statistical power to
detect small changes at the population level. Bias may be introduced if the measurement error of the
outcome is highly correlated with the exposure, but there is no evidence to suggest that individuals with
higher BLLs test systematically lower than their true IQ.
IQ = intelligence quotient.
Note: Two distributions of intelligence test scores. (Left): Based on a mean of 100 (the standardized average, with SD of 15).
(Right): Demonstrating a 5% reduction model, based on a mean score of 95. This is a conceptual model that assumes that the
incremental C-R between Pb exposure and IQ is similar across the full range of IQ and is not based on actual data. The figure
shows that the effect of a small shift in population mean IQ score may result in a larger proportion of individuals with IQ scores
below 70 and a smaller proportion with IQ scores above 130.
Source: Reproduced with permission of Elsevier; from Weiss (19881.
Figure 3-9 Two distributions of intelligence test scores demonstrating the
consequence in a small shift in the mean score.
3.5.1.6.5 Potentially At-Risk Populations
The 2013 Pb ISA described physiologic factors that influence the internal distribution of Pb (U.S.
EPA, 2013). Blood and bone Pb measurements are influenced to varying degrees by biokinetic processes
including absorption, distribution, metabolism, and excretion. These processes are affected by age,
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genetics, diet, and co-exposures to other metals and chemicals, which are summarized in the Executive
Summary and Integrated Synthesis (https://asscssmcnts.cpa.gov/isa/documcnt/&dcid=359536). In
addition to these physiological factors, several population characteristics that explain differential Pb
exposure have been identified. These factors included age, sex, race and ethnicity, proximity to Pb
sources, and residential sources and are also discussed in the Executive Summary and Integrated
Synthesis (https://asscssmcnts.cpa.go\7isa/documcnt/&dcid=359536). The factors potentially related to
increased risk of Pb-induced cognitive effects (i.e., factors identified in epidemiologic studies that
conduct stratified analyses and compare the magnitude of the observed association across stratum) are
discussed below.
Age
This section on Potentially At-Risk Populations emphasizes stratified results described in some
epidemiologic studies, as opposed to the large body of longitudinal studies following mothers and infants
throughout childhood that comprises the most compelling body of evidence in support of conclusions
regarding childhood as an "at-risk" factor. As noted in previous sections of the document (i.e., Section
3.5.1.1 and 3.5.1.6.1), recent evidence supports the finding from the 2013 Pb ISA that cognitive effects in
young children is the outcome that best substantiated to occur at the lowest exposure levels. Strong
evidence indicates increased risk of Pb-induced neurocognitive effects during several childhood lifestages
throughout gestation, childhood, and into adolescence (see Section 3.5.1.6.3). Moreover, the integrated
synthesis (Section 7.4.2.2) of this document concludes that, "In consideration of the evidence base (e.g.,
stratified and longitudinal analyses) and integrating across disciplines of toxicokinetics, exposure, and
health, there is adequate evidence to conclude that children are an at-risk population."
Sex
Multiple epidemiologic studies included in the 2013 Pb ISA examined Pb-related effects on
cognition separately in males and females. Studies on cognition from the CLS cohort and a study in
Poland reported larger magnitude Pb-associated cognitive effects in males (Jedrvchowski et al.. 2009a;
Ris et al.. 2004; Dietrich et al.. 1987). whereas studies from Australia indicated that females were at
increased risk of Pb-associated cognitive effects (Tong et al.. 2000; Baghurst et al.. 1992; Mcmichael et
al.. 1992). While toxicological evidence supporting sex-specific effects of Pb on cognitive function was
summarized in the previous ISA (Virgolini et al.. 2008; Yang et al.. 2003; Mcgivern et al.. 1991). the
number of studies considering sex as a factor was limited. Recent evidence provides more support for the
sex-biased effects. One study reported a male-specific effect (Anderson et al.. 2016) and several studies
demonstrated female-specific effects (Tartaglione et al.. 2020; Verma and Schneider. 2017; Anderson et
al.. 2012; Betharia and Maher. 2012). The evidence supports a conclusion that there are sex-related
differences in the effects of Pb on cognitive function, yet it remains difficult to parse the exact nature and
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direction of sex-specific effects given the variation in outcomes examined, exposure timing and the
considerable number of studies that only reported data from one sex at a time.
Several recent epidemiologic studies examined sex-stratified associations of Pb exposure with
cognitive effects (Tatsuta et al.. 2020; Zhou et al.. 2020b; Desrochers-Couture et al.. 2018; Taylor et al..
2017). Tatsuta et al. (2020) found a decrement in FSIQ score in association with postnatal BLL (|3 =
-9.880 [95% CI: -2.905, 5.831]) among boys, with a smaller less precise association among girls (|3 =
-4.406 [95% CI: -15.94, 7.129]). Prenatal BLL was associated with a smaller and less precise decrease in
FSIQ score among boys (|3 = --3.683 [95% CI: -10.714, 3.349]) but not among girls (|3 = 1.463 [95% CI:
-2.905, 5.831]). A lower BNT score (with cues) was associated with both pre- and postnatal BLL among
boys in this study. Desrochers-Couture et al. (2018) studied the association between cord, maternal and
childhood (3-4 years old) BLLs with cognitive function (WPPSI-III at age 3-4 years) among preschool
aged children in Canada. An association was observed between cord BLL and performance IQ in boys (|3
= -3.28 [95% CI: -5.31, -1.18] per doubling) that was not present in girls (|3 = 0.16 [95% CI: -1.76,
2.06] per doubling). Although associations were imprecise, Taylor et al. (2017) found an association
between increased maternal BLL and IQ decrements in boys but not in girls enrolled in the ALSPAC
study (e.g., -0.29 [95% CI: -1.02, 0.44] versus 0.73 [95% CI: 0.39, 1.33], respectively on the WISC).
Using models that adjusted for co-exposure to metals (Mn and Cd), Zhou et al. (2020b) found no
association between cord BLL and FSIQ in boys or girls. Further, using models that adjusted for other
chemicals (i.e., DDE, HCB, PCBs, and Mn), Oppenheimer et al. (2022) found no statistical evidence of
an interaction between prenatal Pb exposure and sex.
Maternal Stress
Toxicological studies assessed in the 2013 Pb ISA demonstrated that early life exposure to Pb and
maternal stress can result in dysfunction of the HPA axis (U.S. EPA, 2013). Recent toxicological evidence
provides further support for this interaction between maternal stress and Pb exposure during development.
Anderson et al. (2012) demonstrated that exposure to Pb blunted the positive effects of environmental
enrichment on learning in the Morris water maze paradigm. Interestingly, Cory-Slechta et al. (2012)
reported that contrary to the effect previously reported in females, maternal stress improved the
performance of Pb-exposed male offspring in a repeated performance and learning paradigm compared
with Pb-exposed males without maternal stress. While this study supports the interaction between
maternal stress and Pb exposure, it remains unclear whether stress would positively influence other facets
of cognitive function.
Recent epidemiologic studies examined maternal stress as a modifier of the association between
Pb exposure and neurodevelopment. Y Ortiz et al. (2017) used the CRISYS-R questionnaire, which
assesses negative life events across several domains (i.e., financial, legal, career, relationships,
community and home violence, medical problems, other home issues, discrimination or prejudice, and
difficulty with authority) to examine this effect. Third trimester maternal BLL was associated with the
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cognitive component of the BSID in this study and a weak interaction (i.e., lower cognitive scores as BLL
and stress increase) between log-transformed maternal blood Pb and stress was observed (|3 = 1.02 [95%
CI: -0.78, 2.82]). In another study, Zhou et al. (2017) assessed mother-child pairs from the Shanghai
Stress Birth Cohort. Maternal whole blood and maternal prenatal stress levels were assessed at 28-36
weeks of gestation, and the GDS adapted for a Chinese population were administered to children 24-36
months old in the study. No association between prenatal maternal BLL and child cognitive development
was observed; however, an interaction effect was observed such that high maternal stress appeared to
exacerbate the effect of prenatal Pb exposure in several domains, including language (|3 = —33.82 [95%
CI: -60.04, -7.59] per log-10 transformed unit of BLL), while low maternal stress did not (|3 = -1.76
[95% CI: -13.03, 9.51] per log-10 transformed unit of BLL, p-interaction = 0.02).
Other Metal Exposure (Cd, Mn, Hg, As)
A limited number of studies included in the 2013 Pb ISA examined the modification of
association between Pb exposure and cognitive function by other metals (U.S. EPA, 2013). Larger Pb-
associated decrements in IQ (Kim et al., 2009) and neurodevelopment (Henn et al„ 2012) were observed
in children with higher Mn levels. Henn et al. (2012) also observed an interaction between the highest
quintile of Mn and BLL at 12 months (Figure 3-10).
Several recent epidemiologic studies of the association of Pb exposure with FSIQ examined
interactions between Pb exposure and other metals or modification of the Pb-FSIQ association by other
metals. For example, some cross-sectional analyses found evidence that coexposure to Mn may heighten
the effect of Pb in some populations (Martin et al., 2021; Menezes-Filho et al„ 2018), while another study
found no interaction between Pb exposure and Mn (or ALAD) (Lucchini et al., 2012). Several studies of
the association between Pb exposure and infant development also point to possible interactions with other
metals. Lin et al. (2013) observed an interaction with Mn such that children who were highly exposed to
both Mn and Pb had larger neurodevelopmental deficits compared with those with low exposure to just
one or both these metals. Kim et al. (2013b, 2013c) observed a larger decrement in MDI in association
with late pregnancy maternal BLL among those with Cd levels above the median (|3 = -3.20 [95% CI:
-5.35, -1.06]) compared with the decrement among those with Cd levels below the median (|3 = -0.29
[95% CI: -2.88, 2.30]). In contrast to the findings of Henn et al. (2012). Valeri et al. (2017) observed an
association between increasing cord Pb level and neurodevelopmental decrements in children with lower
cord blood Mn and As (|3 = -0.01 [95% CI: -0.02, 0.00]), but not in the group with higher concentrations
of these metals (or metalloids) (|3 = 0.01 [95% CI: -0.05, 0.07]). Nvanzaet al. (2021) also observed
modification of maternal blood Pb and global neurodevelopmental status by blood Hg concentrations.
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BPb = blood Pb; MDI = Mental Developmental Index; Mn = manganese.
Source: Henn et al. (20121.
Figure 3-10 Scatter plots and regression lines of blood Pb level and 18-month
Mental Developmental Index among children in manganese (A)
quintiles 1-4 and (B) quintile 5.
Only one recent animal study incorporated combined exposure to Pb and Mn at relevant levels
(<30 (ig/dL Pb). Betharia and Maher (2012) reported that both Pb and Mn individually impaired memory
in the Morris water maze, but the effects of the mixture were not significantly different from those of the
control. Interestingly, during the learning (acquisition) phase only, the mixture enhanced performance,
suggesting a possible antagonistic effect of these two metals on the development of spatial learning
processes. Given the lack of evidence available on combined metal exposures, the possible interaction
between multiple metals at relevant levels in animals remains unclear.
Many recent toxicological studies provided evidence for the interaction of Pb and other metals
(Mn, Cd, Ar, Hg, Fe) but were not PECOS-relevant (e.g., in vitro studies, high levels, non-mammalian
models) and are summarized in the biological plausibility Section 3.3
Socioeconomic Status
SES has been examined as an effect modifier in multiple studies of Pb-induced cognitive effects
(U.S. EPA, 2013, 2006). Larger blood Pb-associated decreases in cognitive function were found with
lower SES in several studies (Ris et al., 2004; Tong et al„ 2000; Bellinger et al., 1990). In contrast, a
meta-analysis of eight studies found a smaller decrement in FSIQ for studies in disadvantaged
populations than for studies in advantaged populations (Schwartz, 1994a). While the results indicate that
BLL is associated with FSIQ deficits in both higher and lower sociodemographic groups, they do not
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clearly indicate whether groups with different SES differ in Pb-related changes for cognitive function
(Murphy et al., 2013).
No recent epidemiologic studies examined SES as a modifier of the association between Pb
exposure and cognitive effects in children.
Race/Ethnicity
The evidence reviewed in the 2013 Pb ISA pertaining to the modification of the effect of Pb
exposure on cognitive function in children by race or ethnicity was limited to one study (U.S. EPA,
2013). Miranda et al. (2007) presented data indicating that the association between early childhood
exposure to Pb and declines in reading and mathematics scores was similar between Black and white
children. In a recent study, Braun et al. (2018) examined the effects of residential exposure interventions
on dust Pb loadings, BLL, and neurodevelopmental outcomes in children (4 to 8 years old). Although no
intervention effect on BLL was found, overall, the geometric mean childhood BLLs for children 1 to 8
years old was lower in non-Hispanic Black children (See Appendix 2: https://assessments.epa.gov/isa/
document/&deid=359536). No intervention effect on other neurodevelopmental outcomes (e.g., FSIQ,
BSID, BRIEF) were observed.
Pre-existing Disease
Studies that examined the effect of Pb exposure on cognitive function in children across strata
defined by pre-existing disease status were not reviewed in previous assessments (U.S. EPA, 2013,
2006).
A recent study examined the association of Pb exposure with IQ and executive functioning using
BRIEF among children with CKD (Ruebner et al., 2019). Concurrent BLL assessment was associated
with FSIQ decrement in adjusted models (|3 = -2.1 [95% CI: -3.9, -0.2] per 1 (ig/dL increase in BLL).
Associations between BLL and behavioral symptoms indicating executive function problems did not
persist in models that controlled for potential confounders including race, poverty, maternal education,
and clinical factors related to CKD.
Nutritional Factors
The 2006 Pb AQCD included studies that indicated individuals with Fe deficiency and
malnourishment had greater inverse associations between Pb and cognition (U.S. EPA, 2006);
nutritional factors were not examined as effect modifiers of the association between Pb exposure and
cognitive effects in children in more recent studies reviewed in the 2013 Pb ISA (U.S. EPA, 2013).
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No recent epidemiologic studies were available to inform this topic. Recent toxicological studies
(that were PECOS-relevant) investigated the influence of different dietary factors on the effects of Pb. Liu
et al. (2022c) exposed rats to 0.2% Pb in drinking water in combination with either standard rodent chow
or a high-fat diet and then assessed cognitive function using the Morris water maze paradigm. High-fat
diet increased the BLL of rats compared with Pb-exposed rats maintained on standard chow. High-fat diet
enhanced the effect of Pb on learning during the acquisition phase compared with the Pb-control diet
group. During the probe trial, Pb-exposed animals (both diets) had significantly fewer crossings into the
target quadrant compared with untreated animals. Interestingly, a similar memory impairment was
observed in the non-Pb + high-fat diet group, suggesting a role for high-fat diets in cognitive impairment
independent of Pb exposure. The contribution of Pb versus high-fat diet remains unclear based on this one
study.
Al-Qahtani et al. (2022) supplemented Pb exposure in mice with green tea extract and reported
that green tea ameliorated the negative effects of Pb exposure on both learning and memory assessed in
an active avoidance paradigm. Additionally, Long et al. (2022) reported that probiotic supplementation
(Limosilactobacilitis fermentum) in Pb-exposed rats partially mitigated the cognitive deficits observed in
an active avoidance paradigm. These studies support a role for dietary factors in the neurotoxicity of Pb
but the diversity of nutritional factors investigated and the small number of studies make it difficult to
determine their importance.
Genetics
Polymorphisms in certain genes have been implicated in the absorption, retention, and
toxicokinetics of Pb in humans (U.S. EPA, 2013, 2006). Studies assessed in the 2013 Pb ISA indicated
that the presence of ALAD variants was associated with an increase in Pb-related cognitive effects in
adults, but there was limited information for children. In studies of children, inverse associations with
poorer rule learning and reversal, spatial span, and planning were exacerbated among those lacking the
DRD4 gene (Froehlich et al., 2007). Two additional studies found no evidence that the
methylenetetrahydrofolate reductase 677T allele or variants of the DRD2 or dopamine transporter
(DAT1) genes modified the effect of Pb on neurodevelopment (Kordas et al„ 2011; Pilsner et al„ 2010).
Several recent studies add to the limited body of evidence in children. Bah et al. (2022) found that
the effect of low Pb exposure on children's IQ was less among those with the ALAD1 genotype. Rooney
et al. (2018) found interaction effects between variants of glutamate ionotropic receptor NMDA-type
subunits 2A and 2B (GRIN2A and GRIN2B) and Pb exposure on performance on tests of learning,
memory, and executive function, which were more pronounced in boys. Kordas et al. (2011) found that
children with the DRD2 TT genotype (variant) scored higher than children with CC genotype (wild type)
on the Bayley MDI and McCarthy memory scale. However, the variants did not modify the relationship
between BLLs and MDI or McCarthy memory scale scores. Bozack et al. (2021) found that prenatal Pb
exposure was associated with DNA methylation in regions annotated to genes involved in
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neurodevelopment. Overall, the evidence pertaining to interactions between genes and Pb exposure in
children remains limited.
Other Factors
No studies that examined maternal smoking as a modifier of the association between Pb exposure
and cognitive effects were included in previous assessments (U.S. EPA, 2013, 2006). Recent studies did
not examine maternal smoking as a modifier of the association between Pb exposure and cognitive
effects in children.
BMI was not examined as an effect modifier of the association between Pb exposure and
cognitive effects in children in studies reviewed in the 2013 Pb ISA (U.S. EPA, 2013). No recent studies
have examined this factor as an effect modifier.
Maternal self-esteem modified the association between BLL and infant development in Surkan et
al. (2008), which was assessed in the 2013 Pb ISA. No recent epidemiologic studies were available to
inform this topic.
Cognitive reserve was not examined as an effect modifier of the association between Pb exposure
and cognitive effects in children in studies reviewed in the 2013 Pb ISA (U.S. EPA, 2013). No recent
epidemiologic studies were available to inform this topic.
3.5.1.7 Summary and Causality Determination: Cognitive Effects in Children
The evidence from epidemiologic and experimental evidence that supports the causality
determination for cognitive effects in children is outlined in Table 3-2. Overall, recent evidence
supports the conclusion from the 2013 Pb ISA that there is a causal relationship between Pb
exposure and cognitive effects in children.
Studies evaluated in the 2013 Pb ISA found a consistent pattern of associations between higher
BLLs and lower FSIQ in children aged 4-17 years (see Figure 4-2 and Table 4-3 (U.S. EPA, 2013)). The
strongest evidence was provided by prospective studies with analyses of the association of blood Pb
measured in early childhood before FSIQ was assessed or with tooth Pb levels typically measured in
dentin reflecting prenatal or early childhood Pb exposure. These prospective studies typically considered
potential confounding by maternal IQ and education, SES, birth weight, smoking exposure, parental
caregiving quality, and in a few cases, other birth outcomes and nutritional factors. Associations were
found in diverse populations (e.g., Boston, MA; Cincinnati, OH; Rochester, NY; Cleveland, OH; Mexico
City, Mexico; Port Pirie, Australia; and Kosovo, Yugoslavia) in studies that examined children recruited
from prenatal clinics, hospital maternity departments, or schools. Studies generally reported high follow-
up participation supported by evidence that selection bias did not explain the associations observed.
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Multiple recent longitudinal studies of children with mean BLLs <5 (ig/dL add to the evidence
informing the relationship between BLL and IQ in children. Heterogeneity in the magnitude and direction
of the associations was present across these studies, however. In a study of Canadian preschool children
with low blood Pb levels, an association between cord blood Pb level and FSIQ was observed, while the
association of childhood concurrent BLLs with FSIQ effectively null (Desrochers-Couture et al., 2018).
In addition, associations were observed in boys but not in girls in several studies (Tatsuta et al., 2020;
Desrochers-Couture et al., 2018; Taylor et al., 2017). There was also some indication that the
heterogeneity across studies could be explained by modeling choices such as confounder adjustment for
other metals. For example, cross-sectional analyses found evidence that exposure to Mn may modify the
association between Pb exposure and IQ in some populations Martin et al. (2021); (Menezes-Filho et al.,
2018). However, studies that adjusted for multiple metals (e.g., Mn, Hg, Cd, and Pb) in regression
models, without examining the interaction between metals, found little evidence of an association
between cord or postnatal BLL and IQ (Zhou et al„ 2020b; Liu et al., 2015), imprecise associations only
in boys (Tatsuta et al., 2020; Desrochers-Couture et al., 2018), or large IQ decrements after adjustment
for Mn, Hg, and ADHD rating score Hong et al. (2015). Overall, recent studies generally corroborated the
epidemiologic observations of associations between Pb exposure and IQ in children with relatively low
blood Pb concentrations (<5 (ig/dL) among some groups of children (see Section 3.5.1.1). Consistent with
findings from the 2013 Pb ISA, studies continue to report associations with prenatal BLL (maternal and
cord blood Pb) and postnatal BLLs measured at various childhood lifestages despite the aforementioned
heterogeneity at BLLs <5 (ig/dL. Overall, the heterogeneity did not weaken the larger body of supporting
evidence.
In the review of the MDI evidence in the 2013 Pb ISA, emphasis was placed on results from
examinations at ages 2-3 years, which incorporate test items more similar to those in school-age IQ tests.
Among these studies, several included children with mean BLLs less than 5 (ig/dL (Henn et al„ 2012;
Jedrychowski et al„ 2009b; Hu et al„ 2006; Bellinger et al„ 1987). Most of the prospective studies
reviewed in previous IS As (U.S. EPA, 2013, 2006) found associations of higher prenatal (cord and
maternal BLL), earlier infancy, and concurrent BLL with lower MDI scores in children aged 2 to 3 years
(Figure 3-10). These blood Pb-associated decrements in MDI were observed in populations with mean
BLLs of 1.3 to 7.1 (ig/dL. Studies typically recruited participants before or at birth without consideration
of Pb exposure or maternal IQ and reported high to moderate follow-up participation as well as
nondifferential loss-to-follow-up. Most studies adjusted for birth outcomes, maternal IQ, and education.
Cord BLLs were associated with MDI, with additional adjustment for SES and HOME score in the
Boston cohort (Bellinger et al., 1987) and for HOME score in the Yugoslavia cohort (Wasserman et al.,
1992). Some studies found a stronger association of MDI with prenatal than child postnatal BLLs ((Hu et
al„ 2006; Gomaa et al., 2002; Bellinger et al., 1987).
Recent studies continue to support associations between Pb exposure measured during prenatal or
childhood lifestages and poorer performance on tests of neurodevelopment, among mothers and infants
with mean BLLs <5 (ig/dL (i.e., maternal (Y Ortiz et al„ 2017; Vigeh et al„ 2014; Kim et al„ 2013b, c),
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cord (Valeri et al., 2017), and postnatal (Lin et al.. 2013) BLLs). Although Zhou et al. (2017) found no
association overall, this study reported decrements on several domains of the GDS among infants of
mothers reporting high maternal stress. Similarly, Y Ortiz et al. (2017) found some evidence of
interactions between Pb exposure and maternal stress. Several studies found interactions between Pb and
Mn or Mn and As (Valeri et al„ 2017; Lin et al„ 2013; Henn et al„ 2012) or Cd exposure (Kim et al„
2013b); Kim et al. (2013c). Overall, recent studies support findings from the previous reviews and extend
the evidence pertaining to modification of the association between Pb exposure and infant
neurodevelopment by maternal stress and exposure to other metals. The MDI and other tests that measure
neurodevelopment in infants and toddlers are not intelligence tests. Notably, MDI scores, particularly
before ages 2-3 years, are not necessarily strongly correlated with later measurements of FSIQ in children
with normal development and thus, are not weighted heavily in the consideration of causality (U.S. EPA,
2013).
Experimental animal studies evaluated in the 2013 Pb ISA demonstrated that prenatal and early
postnatal or lifetime Pb exposure alters brain development via changes in synaptic architecture and
neuronal outgrowth, leading to impairments in memory and learning (Sections 4.3.10.4, 4.3.10.10, and
4.3.2.3 of U.S. EPA (2013)). A small number of recent experimental animal studies were designed to
compare exposures across multiple different developmental windows (Barkur and Bairy, 2015b; Xiao et
al„ 2014); these studies reported similar magnitudes of effects between developmental windows,
suggesting that individual periods of development may be similarly sensitive to Pb. Generally, longer
exposures that spanned multiple developmental periods (e.g., preconception through lactation) produced
not only the highest BLLs but the largest effects on cognition (Zhou et al., 2020a; Zhu et al„ 2019b).
Overall, these studies provide strong support for observations in epidemiologic studies that Pb exposure
during the prenatal, childhood, and adolescent lifestages is associated with cognitive effects. Recent
animal studies also provide evidence to support the observation that development (i.e., preconception,
gestation, lactation) may represent a critical window for Pb exposure to cause cognitive impairment later
in life (see Section 3.6.1). In rodents, developmental exposure to Pb was consistently associated with
persistent cognitive effects observed both early (Tartaglionc et al., 2020; Zhao et al., 2018; Barkur and
Bairy, 2015b; Anderson et al„ 2012) and later in life (Liu et al„ 2022c; Xiao et al„ 2014; Betharia and
Maher. 2012).
Learning, memory, and executive function are domains of cognitive function that are related to
intelligence, and several are evaluated in the subtests of FSIQ. Additionally, indices of memory, learning,
and executive function are comparable to endpoints examined in experimental animal studies. The studies
evaluated in the 2006 Pb AQCD and 2013 Pb ISA did not clearly indicate associations between higher
BLL and poorer performance on neuropsychological tests of memory or learning (U.S. EPA, 2013). The
ascertainment of the outcomes varied across studies, potentially explaining the heterogeneity of the
epidemiologic observations. Notably, evidence for both memory and learning decrements from
prospective analyses of several established cohorts (i.e., Rochester, Boston, and Cincinnati) was mixed
(Canfield et al„ 2004; Ris et al., 2004; Stiles and Bellinger, 1993; Bellinger et al„ 1991; Dietrich et al.,
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1991). Cross-sectional studies included in the previous ISA, however, generally found associations
between higher concurrent BLLs and poorer learning and memory. Several recent studies of children with
mean BLLs <5 (ig/dL add to the evidence informing the association of Pb exposure with performance on
tests of memory and learning; however, these recent studies do not enhance the consistency of the
evidence as a whole. Some of the available studies consider co-exposure to other chemicals and metals as
confounders (Tatsuta et al., 2014) despite evidence that such co-exposures may interact with or modify
the association between Pb and the outcomes (Yorifuji et al„ 2011). Several recent studies of rodents with
exposure resulting in mean BLLs <30 (ig/dL add to the evidence indicating coherence between the
epidemiologic and toxicological findings pertaining to learning and memory observed in the 2013 Pb
ISA.
Strong evidence of associations between Pb exposure and indices of executive function was
described in the 2013 Pb ISA. Studies included prospective analyses of several birth cohorts with
moderate to high follow-up rates in Boston and Rochester that examined BLLs before the outcome
assessment and adjusted for several potential confounding factors (Canfield et al., 2004; Canfield et al.,
2003b; Bellinger et al., 1994a; Stiles and Bellinger, 1993). Recent studies relying on parent or teacher
behavioral ratings on BRIEF did not generally report associations. The previous ISA did not incorporate
any evidence of the relationship between Pb exposure and executive function in animal models. Recent
studies from a single laboratory provided evidence that Pb exposure broadly impairs measures of
executive function in a reversal learning paradigm. These effects were sex-specific, with greater effects
reported in males. While these reports are consistent with one another, evidence for the association
between Pb exposure and impaired executive function in animal models with BLLs <30 (ig/dL remains
limited.
As described in Sections 3.5.1.1 and 3.5.1.2 and summarized above, heterogeneity in the
epidemiologic results for FSIQ and infant development at BLLs <5 (ig/dL may be explained in part by
sex, exposure to other metals, or maternal stress. Experimental animal studies offer some support for the
observations regarding sex and maternal stress. The limited evidence evaluated in the 2013 Pb ISA
(Virgolini et al., 2008; Yang et al„ 2003; Mcgivern et al., 1991), combined with recent evidence, provides
more consistent support for the sex-biased effects in both male (Anderson et al„ 2016) and female
(Tartaglione et al., 2020; Verma and Schneider, 2017; Anderson et al„ 2012; Betharia and Maher, 2012)
animals. The exact nature and direction of sex-specific effects given the variation in outcomes examined
remains unclear, however. Toxicological studies assessed in the 2013 Pb ISA demonstrated the potential
for Pb and maternal stress to result in dysfunction of the HPA axis (U.S. EPA, 2013). Recent
toxicological evidence provides further support for the interaction between maternal stress and Pb
exposure during the exposure period (Anderson et al., 2012; Cory-Slechta et al., 2012) but it remains
unclear whether stress would positively influence some facets of cognitive function. Given the lack of
evidence available on combined metal exposures, the possible interaction between multiple metals at
relevant levels in animals remains unclear.
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Poorer academic performance and achievement is linked with lower FSIQ and may have
important implications for success later in life (U.S. EPA, 2013). In children aged 5 to 18 years higher
blood Pb measured at various lifestages, including early childhood, and tooth Pb levels, which were
generally measured in dentin and reflect Pb exposure during the prenatal or early childhood period, were
associated with poorer performance on tests of math, reading, and spelling skills, lower probability of
high school completion and lower-class rank, and lower teacher ratings of academic functioning (U.S.
EPA, 2013). Recent studies in populations of children (age 6-16 years) enrolled in school districts
including North Carolina, Detroit, and Chicago with BLLs <5 (ig/dL support and extend these
observations of poorer academic performance in association with increasing Pb exposure in populations
with mean BLLs <5 (ig/dL.
Recent studies support and extend the evidence pertaining to the effect of Pb exposure on
cognitive function in children at low BLLs. Compelling evidence for a larger decrement in cognitive
function per unit increase in blood Pb among children with lower mean blood Pb concentrations,
compared with children with higher mean blood Pb concentrations, was supported by a reanalysis of a
pooled international dataset (Crump et al., 2013). The larger incremental effect of Pb on cognitive effects
at the lower (relative to higher) end of the study population Pb exposure distribution has been observed
across biomarkers [i.e., bone (Wasserman et al„ 2003; Wasserman, 2003), plasma (Hu et al„ 2006) and
blood (Section 4.3.12, Figure 4-15, and Table 4-16 of the 2013 Pb ISA U.S. EPA (2013))! and across
different cognitive function endpoints [i.e., infant development (Hu et al„ 2006), IQ (Wasserman et al„
2003; Wasserman, 2003) and academic achievement (Evens et al.. 2015).! Crump et al. (2013) also
supported the finding of Lanphear et al. (2019) that the C-R function in the pooled analysis of blood Pb
level and IQ is adequately modeled as linear at BLLs <10 (ig/dL. Recent studies with an adequate range
of Pb exposure measured during relevant time periods that would be required to further evaluate
exposure-response relationships were limited. Considering the collective body of studies, no evidence of a
threshold for cognitive effects in children across the range of BLLs examined in epidemiologic studies
was reported.
The total body of evidence evaluated is sufficient to conclude that there is a causal
relationship between Pb exposure and decrements in cognitive function in children. This causality
determination is the same as the conclusion in the 2013 Pb ISA, reflecting the consistency of the results
from epidemiologic studies of FSIQ, Bayley MDI, and academic performance and achievement, as well
as the coherence of evidence across epidemiologic and toxicological studies of learning and memory. The
pattern of associations consistently observed for tooth Pb levels and blood Pb levels measured at various
lifestages or time periods reduces uncertainty regarding the temporal association observed in cross-
sectional analyses of concurrent blood Pb levels with cognitive decrements in children. Notably
concurrently measured tooth Pb levels in dentin generally reflect the prenatal or childhood Pb exposure
that precedes the assessment of the outcome. Biological plausibility is provided by studies that describe
pathways involving the interaction of Pb with cellular proteins, in some cases competing with and
displacing other biologically relevant cations. This interaction leads to increased oxidative stress and the
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presence of inflammation, which can have widespread effects on brain structure and function, as well as
disruptions of Ca2+ signaling. These disruptions can result in altered brain signaling and contribute to the
development of neurological health effects. Recent studies support the conclusion of the 2013 Pb ISA that
Pb-associated cognitive effects in children occur in populations with mean BLLs between 2 and 8 (ig/dL.
As noted in the 2013 Pb ISA, this conclusion was based on studies that examined early childhood BLLs
(i.e., age <3 years), considered peak BLLs in their analysis (i.e., peak <10 (.ig/dL). or examined concurrent
BLLs in young children (i.e., age 4 years). One recent study of Canadian preschool aged children from
mainly middle- to upper-middle SES families with low Pb exposure (mean concurrent blood Pb level
0.70) did not find an association between concurrent Pb exposure and performance on WPPSI at age 3 to
4 years (Desrochers-Couture et al.. 2018). Other recent studies found associations of Pb exposure with
cognitive effects in children with mean BLLs <2 (ig/dL; however, the studies with mean BLLs <2 (ig/dL
lack the aforementioned attributes (i.e., early childhood BLLs, consideration of peak BLLs, or
examination of concurrent BLLs in young children) and exhibit heterogeneity in both the magnitude and
precision of the associations at the lowest blood Pb concentrations. The observed heterogeneity may be
explained in part by the underlying distribution and complex relationship between covariates in the
populations studied, including sex, maternal stress, and co-exposures to other metals and neurotoxic
chemicals, at relatively low BLLs (<5 (.ig/dL). Overall, epidemiologic and toxicological studies continue
to strongly support the finding that exposure during multiple lifestages (prenatal through adolescence and
early adulthood) is associated with cognitive effects in children. No evidence of a threshold for cognitive
effects in children across the range of BLLs examined in epidemiologic studies was reported.
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Table 3-2 Summary of evidence Indicating a causal relationship between Pb exposure and cognitive effects
in children
Rationale for C^sality Key Evidence13 References'3 Levels Associated
Determination3 wjth EffectsC
Consistent associations from multiple, Evidence from prospective studies for decrements U.S. EPA (2013) Blood Pb (various
prospective epidemiologic studies with in FSIQ in association with prenatal, earlier Section 4 3 2 1 Table 4-3 t'me Peri°ds and
relevant BLLs childhood, peak, concurrent, lifetime average BLLs ' lifestages): Means 3-
and tooth Pb levels in children ages 4-17 yr in 16 |jg/dL
multiple U.S. locations, Mexico, Europe, Australia.
Recent prospective studies observe associations of Section 3.5.1.1 Blood Pb (various
Pb with FSIQ; however, heterogeneity in the time periods and
magnitude and direction of the associations is lifestages) <5 |jg/dL
present. (<2 |jg/dL in some
studies)
Some recent epidemiologic studies indicate Section 3.5.1.6.5
potential effect modification or interactions of Pb
with sex, other metals, and maternal stress
potentially explaining heterogeneity in the observed
associations at BLLs <5 |jg/dL.
Evidence from prospective studies for lower scores
on tests of executive function and academic
performance in association with earlier childhood or
lifetime average BLLs or tooth Pb levels in children
ages 5-20 yr in multiple U.S. locations, U.K., New
Zealand.
U.S. EPA (2013)
Blood Pb (various
time periods and
lifestages) <5 |jg/dL
Recent evidence generally relies on outcome Section 3.5.1.4
ascertainment based on the BRIEF is inconsistent.
The direction and magnitude of associations were U.S. EPA (2013) and Section 3.5.1.3
not consistent for learning and memory. Recent
evidence does not enhance the consistency.
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Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker
Levels Associated
with Effects0
Supporting evidence from cross-sectional studies of
children ages 3-16 yr, but most did not consider
potential confounding by parental caregiving
quality. Includes large NHANES III analysis.
Several studies indicate supralinear C-R
relationship, with larger decrements in cognitive
function per unit increase in blood Pb at lower BLLs
in children ages 5-10 yr. Reanalysis of international
pooled analysis substantiates this finding.
Epidemiologic evidence helps rule out
chance, bias, and confounding with
reasonable confidence
Several epidemiologic studies found associations
with adjustment for SES, maternal IQ and
education, HOME score. Several adjust for birth
weight, smoking. A few, nutritional factors.
U.S. EPA (2013)
Section 3.5.1.6.2
Experimental animal studies with
relevant exposures provide coherence
and help rule out chance, bias, and
confounding with reasonable
confidence
Impaired learning and associative ability in juvenile
and adult animals as indicated by performance in
tasks of visual discrimination, water maze, y maze,
and operant conditioning with schedules of
reinforcement with relevant dietary Pb exposure.
U.S. EPA (2013)
Section 4.3.2.3
Blood Pb (after
prenatal/ lactation,
lactation only,
prenatal/lifetime Pb
exposure): 10-25
pg/dL
Recent studies of executive function in rodents add
to the evidence; however, evidence for impaired
executive function in animal models with BLLs <30
|jg/dL remains limited.
Section 3.5.1.4.2
Experimental animal studies with
relevant exposures provide coherence
for epidemiologic observations of effect
modification by sex or interactions of Pb
with other metals or maternal stress
Recent studies in rodents suggest that factors such
as sex and maternal stress may influence the
effects of Pb on cognitive function.
Section 3.5.1.6.5
Biological plausibility demonstrated
Pathways involving oxidative stress, inflammation
and Ca2+ signaling result in impaired neuron
development, synaptic changes, LTP, and
neurotransmitter changes.
U.S. EPA (2013)
Section 3.6
U.S. EPA (2013) Blood Pb (various
time periods and
lifestages) <5 |jg/dL
U.S. EPA (2013)
Section 3.5.1.6.1
Crump et al. (2013)
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Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker
Levels Associated
with Effects0
Recent studies support and extend findings related Section 3.3
to overt nervous system effects.
BLL = blood lead level; BRIEF = Behavior Rating Inventory of Executive Functions; Ca2+ = calcium ion; C-R = concentration-response; FSIQ = full-scale intelligence quotient; HOME
= Health Outcomes and Measures of the Environment; IQ = intelligence quotient; LTP = long-term potentiation; NHANES = National Health and Nutrition Examination Survey; Pb =
lead; SES = socioeconomic status; yr = year(s).
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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3.5.2 Externalizing Behaviors: Attention, Impulsivity, and Hyperactivity in
Children
The evidence evaluated in the 2013 Pb ISA was sufficient to conclude that there is a "causal
relationship" between Pb exposure and effects on attention, impulsivity, and hyperactivity in children.
Several prospective studies demonstrated associations of blood Pb measured years before outcomes or
tooth Pb levels, which were generally measured in dentin and reflect prenatal and early childhood Pb
exposure, with attention decrements and hyperactivity in children (7-20 years) as assessed using objective
neuropsychological tests and rated by parents and teachers. Most of the prospective studies examined
representative populations with no indication of participation that was conditional on BLLs and behavior.
The results from the prospective studies were generally adjusted for potential confounding by SES as well
as parental education and caregiving quality, with some studies also considering parental cognitive
function, birth outcomes, substance abuse, and nutritional factors. With respect to the timing of exposure
discerned from prospective studies, blood Pb-associated attention decrements and hyperactivity were
found in populations with prenatal (maternal or cord) or postnatal (i.e., 3-60-month average, age 6 years,
or lifetime average through age 11-13 years) mean BLLs of 7 to 14 (ig/dL and in groups with BLLs >10
(ig/dL at 30 months of age. Biological plausibility for these observations in children was provided by
experimental animal studies that demonstrated increases in impulsivity or impaired response inhibition
with relevant postweaning and lifetime Pb exposures that resulted in BLLs of 11 to 30 (ig/dL.
Demonstrated Pb-induced impairments in neurogenesis, synaptic pruning, and dopamine transmission in
the prefrontal cerebral cortex, cerebellum, and hippocampus also supported the biological plausibility of
the associations observed in the epidemiologic studies. Although coherence across and within lines of
evidence was demonstrated, the small number of studies of diagnosed ADHD were limited by their cross-
sectional or case-control design, inconsistent adjustment for SES and parental education, and lack of
consideration for potential confounding by parental caregiving quality.
There are three major domains of externalizing behavior disorders: (1) ADHD, (2)
undersocialized aggressive conduct disorder, and (3) socialized aggressive conduct disorder (as reviewed
in (Whitcomb and Merrell. 2012)). Although these domains are interrelated, to the extent possible, this
Section (3.5.2) will maintain a similar structure as the 2013 Pb ISA by focusing on the ADHD domain,
which encompasses characteristics including but not limited to short attention span, distractibility,
impulsivity, and hyperactivity. Within the ADHD domain of externalizing behaviors, most epidemiologic
studies of Pb exposure focus on attention, impulsivity, and hyperactivity. Some epidemiologic studies
examined composite indices of multiple behaviors, and a few studies have examined physician-diagnosed
ADHD. Domain-specific neuropsychological assessments of attention, impulsivity, and hyperactivity
with strong psychometric properties and rigorous validation were emphasized in the 2013 Pb ISA and
provide the strongest evidence for the causality determination. Studies that evaluated the association of Pb
exposure with externalizing behaviors assessed using teacher and parent ratings, which are generally
reliable and valid instruments that predict functionally important outcomes, contributed to the overall
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evidence (Desrochers-Couture et al.. 2019; Fruh et al., 2019; Nigg et al.. 2016; Hong et al., 2015;
Gittleman and Eskenazi, 1983).
Control for confounding is considered an attribute of a well-conducted, high-quality study.
Greater weight is given to studies that consider important potential confounders in their design or
statistical analyses. As noted in the 2013 Pb ISA (U.S. EPA, 2013), associations between Pb biomarker
levels and externalizing behaviors may be confounded by parental SES, education, and IQ; nutritional
status; and the quality and stability of the caregiving environment (often evaluated using the HOME score
(Totsika and Svlva. 2004)). The research available for evaluation in the 2013 Pb ISA did not establish a
direct relationship between parental psychopathology and child Pb exposure or one between parental
psychopathology and poorer parental caregiving quality. Thus, parental psychopathology itself was not
considered to be a potential confounder of associations between child Pb and externalizing behaviors.
Although parental psychopathology was hypothesized to modify the association between Pb exposure and
externalizing behavior, no studies available for evaluation in the 2013 Pb ISA evaluated parental
psychopathology as an effect modifier. To the extent that parental psychopathology could affect child Pb
exposure indirectly through parental caregiving quality, however, it was noted that confounder control
would be achieved in studies that included an adjustment for the HOME score or similar metrics.
Greater emphasis is also placed on prospective studies with repeated assessments of BLLs and
studies of children with BLLs that are less influenced by higher past Pb exposures (i.e., younger children).
Studies assessing effects in populations with BLLs that are most relevant to current U.S. children (e.g., <5
(ig/dL) are also emphasized, e.g., (Cho et al„ 2010; Nicolescu et al., 2010; Chandramouli et al„ 2009;
Nigg et al„ 2008; Chen et al„ 2007; Chiodo et al., 2007). In the current ISA, when considering the causal
relationship of Pb exposure with attention, impulsivity, and hyperactivity, PECOS statements (see Section
3.2) were refined to focus on the most informative studies. Longitudinal epidemiologic studies with mean
(or central tendency) BLLs <5 (ig/dL are highlighted in the text as are the most reliable biomarkers of Pb
exposure (i.e., blood, bone, teeth, or nails). Consideration of potential confounding and modification of
the observed associations by the aforementioned factors was evaluated when considering the overall
quality of the study. Measures of central tendency for Pb biomarker levels used in each study, along with
other study-specific details, including study population characteristics and select effect estimates, are
highlighted in evidence inventory Table 3-7E (Epidemiologic Studies) and Table 3-7T (Toxicological
Studies). In addition, studies with central tendency blood Pb concentrations that exceed 5 (ig/dL are
extracted into Table 3-8E of Section 3.7 (Evidence Inventories). An overview of the recent evidence is
provided below. Overall, recent studies generally support findings from the 2013 Pb ISA.
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3.5.2.1
Attention in Children
3.5.2.1.1 Epidemiologic Studies of Attention in Children
Attention is the ability to maintain a consistent focus on an activity or relevant stimuli and can be
assessed by examining sustained attention, concentration, or distractibility. The preponderance of
evidence pertaining to the externalizing behaviors included in the 2013 Pb ISA evaluated the association
of Pb exposure with measures of attention (U.S. EPA, 2013). Most prospective studies found associations
of blood or tooth Pb levels with decrements in neuropsychological tests of attention as well as parent and
teacher ratings of attention. One strength of the prospective studies is that they characterized the sequence
of Pb exposure (i.e., prenatal blood Pb, postnatal blood Pb before the outcome, concurrent, lifetime
average blood Pb, and tooth Pb [i.e., generally measured in dentin, reflecting prenatal, early childhood or
cumulative Pb exposure depending on the assessment method]), establishing the temporal relationship
between exposure and outcome. In addition, the studies reported moderate to high follow-up participation
that was not conditional on blood or tooth Pb levels and controlled for important confounders (i.e.,
parental education, IQ, and caregiving quality; SES). Overall, these studies showed a pattern of lower
attention with higher blood or tooth Pb level (see Figure 4-9 and Table 4-11 of the 2013 Pb ISA (U.S.
EPA, 2013)). Mean BLLs were generally within the range of 7-14 (ig/dL for most of the prospective
studies, and cross-sectional studies generally supported findings from the longitudinal analyses (see
Section 4.3.3.1 of the 2013 Pb ISA (U.S. EPA, 2013}!
A small number of recent longitudinal epidemiologic studies of children with relatively low BLLs
(i.e., <5 (ig/dL) add to the evidence for associations between Pb exposure and decrements in
neuropsychological tests of attention or parent and teacher ratings of behaviors that indicate attention
problems (see Section 3.5.2.4).
Ncugcbaucr et al. (2015) conducted an analysis of the Duisburg birth cohort data to examine the
association of maternal BLLs at 32 weeks gestation with performance on neuropsychological tests of
attention (Test of Attentional Performance for Children [KiTAP]) and parent-rated ADHD behaviors on
the German Symptom Checklist for ADHD (Fremdbeurteilungsbogen fur
Aufmerksamkeitsdefizit/Hyperaktivitatstorungen [FBB-ADHS]) in childhood. Maternal blood Pb was
most strongly associated with specific KiTAP subtests, i.e., number of omissions (geometric mean ratio
[GMR] = 1.15 [95% CI: 1.00, 1.33]) and reduced performance speed (GMR= 1.14 [95% CI: 0.98, 1.33]).
Maternal blood Pb concentration was positively associated with the inattention component of the FBB-
ADHS indicating that inattention increases with increasing BLL (GMR = 1.05 [95% CI: 0.99, 1.12]).
In a study of a subset (n = 27) of Inuit children (Boucher et al., 2012b) (see Section 3.5.2.5) that
used a modified Posner paradigm to assess the association between prenatal and concurrent childhood Pb
exposure with visuospatial attention, vigilance, and impulsivity, Ethier et al. (2015) found that concurrent
ln-transformed blood Pb was associated with some tests of attention including longer reaction times (|3 =
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0.52 [95% CI: -0.10, 1.14] per SD increase in ln-transformed Pb) in a model adjusted for age, sex, and
current PCB exposure. In another study, Tatsuta et al. (2014) adjusted for PCBs and MeHg (in addition to
maternal IQ and family income) and found no associations with sequential processing score (-2.14 [95%
CI: -12.80, 8.53] per unit of log transformed BLL [base not specified]) or mental processing score (-3.32
[95% CI: -12.4, 5.77] per unit of log transformed BLL [base not specified])). Yorifuii et al. (2011) found
that cord blood Pb was associated with some neuropsychological tests of attention and working memory
on the WISC-R (i.e., digit span) and that the interaction between cord blood Pb and cord Hg level may be
less than additive. For example, the association of cord Pb with performance on the digit span forward at
age 7 was |3 = —0.11 [95% CI: -0.29, -0.07] per log-transformed unit of BLL without accounting for the
interaction between cord Pb and cord Hg concentration. This association was more pronounced when Hg
exposure was lower. Specifically, the unstandardized associations of log-transformed cord BLL with the
neuropsychological test outcomes were most discernable among children with hair Hg concentrations
below 2.61 (ig/g, which was the lowest cord Hg concentration. For example, a lower digit span forward
score on the WISC-R (|3 = -1.70 [95% CI: -3.12, -0.28]) at age 7 and a lower digit span backward score
on the WISC-R (|3 = -2.73 [95% CI: -4.32, -1.14]) at age 14 were observed among children with the
lowest Hg exposure.
Ruebner et al. (2019) evaluated the association between BLLs and attention and hyperactivity
among children with CKD. Attention was assessed using either Couriers' Kiddie Continuous Performance
(K-CPT; 4-5 years) or Couriers' CPT II (>6 years), which produce scores for omission and commission
errors, correct detection rate, response variability, reaction time, and summary measures for sustained
attention and inhibitory control. A 1.8 T-score point increase (i.e., worse performance) in CPT variability
(95% CI: 0.2, 3.5), which indicates problems with sustained attention and attention regulation, was
associated with childhood blood Pb (on average, ~2 years before outcome ascertainment) in this study.
Covariates considered as potential confounders included race, poverty, maternal education, and factors
related to CKD. The median BLL in this study was 1.2 (ig/dL.
A small number of studies examined the gene-environment interaction between Pb exposure and
genotypes associated with attention decrements. Roonev et al. (2018) studied children in Lisbon, Portugal
to determine the association of baseline BLL (8-12 years old) and variants of GRIN2A and GRIN2B,
which regulate neurodevelopmental processes, with performance on neuropsychological tests, including
tests of attention, during the 7-year follow-up period. A pattern of association indicating poorer
performance on tests of attention with increasing baseline Pb exposure was not observed. Choi et al.
(2020) enrolled children (5-18 years old) with ADHD and healthy controls without ADHD to evaluate
interactions between Pb exposure and noradrenergic pathway-related genotypes (i.e., [DAT1], dopamine
receptor D4 [DRD4], and alpha-2A-adrenergic receptor [ADRA2A]). ADHD was assessed using the
ADHD rating scale (ADHD-RS) and neuropsychological tests of attention (i.e., CPT and SCWT) were
also administered. BLLs were associated with omission errors (|3 = 3.75 [95% CI; 0.09, 7.40]) in models
adjusted for IQ, age, and sex, which were found to partly mediate the effect of Pb on ADHD symptoms in
a path analysis model. An interaction effect was detected between the ADRA2A Dral genotype and Pb
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levels on omission errors (|3 = 5.07 [95% CI: 0.20, 9.93]). Multiple comparisons were made in this
analysis (e.g., associations with additional CPT components including commission errors, response time,
and response time variability, ADHD-RS components and SCWT components were not observed),
increasing the likelihood of chance findings.
Summary
Prospective studies in the 2013 Pb ISA showed strong support for an association between pre-
and postnatal Pb exposure (range: 7-14 (ig/dL) and decreased scores on neuropsychological tests and
parent/teacher ratings of attention. Cross-sectional studies from the 2013 Pb ISA corroborated these
observations. A small number of recent studies reported associations of maternal and cord BLLs <5 (ig/dL
with some measures of inattention; however, the results for multiple subtests were reported, potentially
increasing the likelihood of chance findings. Recent studies add to the limited evidence regarding co-
exposure to Hg and gene-environment interactions (see Section 3.5.2.6.3).
3.5.2.1.2 Toxicological Studies of Attention
In support of the associations described in the preceding sections for BLL with attention
decrements in children, studies have found Pb-induced decreases in attention in animals, although results
have not been consistent across studies. Although tests in animals often measure aspects of both attention
and impulsivity, behaviors measured with signal detection tests with distraction can be inferred as
predominately assessing sustained attention. In this test, animals earn food rewards by responding to a
target stimulus and not responding to a distracting light. Poorer sustained attention and greater
distractibility are indicated by lack of response to the target and increased response to the distracter light,
respectively. The 2006 Pb AQCD (U.S. EPA, 2006) reported inconsistent effects of Pb exposure in
animals on performance in this test. For example, postweaning Pb exposure that produced BLLs of 16
and 28 (ig/dL induced small decreases in attention in adult rats, as indicated by small increases in
omission and commission errors but only during sessions with long intervals between stimuli (Brockel
and Cory-Slechta, 1999). Lifetime Pb exposure from birth (mean peak BLLs of 15 and 25 (ig/dL for the
50 and 100 (ig/kg/day groups, respectively) was found to induce distractibility in monkeys at age 9-10
years, as indicated by increased responses to irrelevant cues, i.e., distracting stimuli, in a spatial
discrimination reversal task. Repeated reversal testing revealed that these deficits likely were not due to
sensory or motor impairment (Gilbert and Rice, 1987).
In animals, Pb-induced decrements in attention have been inferred from tests designed to assess
impulsivity but that have elicited behaviors that suggest deficits in attention. For example, a study
reported that impaired performance on auditory threshold tasks in Pb-exposed monkeys was likely due to
lack of attention (Laughlin et al., 2009). Rhesus monkeys were exposed to Pb acetate from gestation
(drinking water of mothers, 3 months prior to mating) to birth or postnatally from birth to age 5.5 months
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at weaning, resulting in bone Pb levels of 7 and 13 |ig/g for prenatal and postnatal groups at 11 years of
age, respectively, and average BLLs of 35 and 46 (ig/dL, respectively, during Pb exposure. Animals were
tested at age 13 years when BLLs had returned to baseline levels. The inability of some of the monkeys to
engage or focus attention on the task at hand yielded fewer available measurements in Pb-exposed
animals versus controls. These observations were made in monkeys with higher peak BLLs than those
relevant to this ISA. No recent studies have evaluated attention in animals following exposure that
resulted in BLLs relevant to the current ISA.
3.5.2.2 Impulsivity in Children
3.5.2.2.1 Epidemiologic Studies of Impulsivity in Children
Measures specific to impulsivity were examined in relatively few epidemiologic studies of
children compared with measures of attention, and most studies including evaluations of impulsivity that
were included in the 2013 Pb ISA were cross-sectional in design (U.S. EPA, 2013). The available
evidence indicated Pb-associated poorer performance on tests of response inhibition. Response inhibition
is a measure of impulsivity and has been assessed in children via stop signal tasks, which measure the
execution of action in response to stimuli and the inhibition of that action when given a stop signal.
Associations of blood and tooth Pb with parent and teacher ratings of impulsivity were also reported.
These studies generally adjusted for potential confounding by SES, sex, parental education, and smoking;
however, parental IQ or caregiving quality was not examined in most studies. The relatively small body
of epidemiologic evidence (see Figure 4-9 and Table 4-11 (U.S. EPA, 2013)) was coherent with results
from experimental animal studies (see Section 4.3.3.1 (U.S. EPA. 2013)).
Analyses of Inuit children have been conducted since the 2013 Pb ISA, examining Pb exposures
and impulsivity. Boucher et al. (2012a) examined response inhibition deficits assessed with the Go/No-
Go task and event-related potentials (ERPs) derived from electroencephalogram (EEG) recordings during
task performance among Inuit school children residing in Arctic Quebec. Cord and concurrent blood Pb
concentrations were associated with increased impulsivity after adjustment for covariates including child
age, sex, SES, maternal nonverbal reasoning abilities, and Hg (Figure 3-11). In addition, Ethier et al.
(2015) studied a subset of this population (n = 27) using a modified Posner paradigm to assess the
association between pre- and concurrent childhood Pb exposure with visuospatial attention, vigilance, and
impulsivity. The study found that cord Pb was associated with greater impulsivity (|3 = 0.42 [95% CI:
0.08, 0.76] per SD increase in ln-transformed Pb).
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0.8-2.5 2.5-3.7 3.7-5.6 5.6-20.9
Cord blood Pb (|jg/dL)
0.8-2.5 2.5-3.7 3.7-5.6 5.6-20.9
Cord blood Pb (|ig/dL)
Source: Boucher et al. (2012a)
Figure 3-11
0.4-1.3 1.3-2.0 2.0-2.9 2.9-12.8
11-year blood Pb (|ig/dL)
Mean ± standard deviation behavior performance in the Go/No-Go
task according to quartiles of exposure for (A and B) cord blood
Pb and (C) childhood blood Pb level at age 11 years.
Summary
Studies of impulsivity in children in the 2013 Pb ISA were limited by their quantity and lack of
temporality but generally indicated associations of Pb exposure with worse scores on tests of response
inhibition and on parent and teacher ratings of impulsivity. These studies also often lacked confounder
control for parental IQ or caregiving quality, which are key potential confounders. Recent analyses of
Inuit children add support for the relationship between Pb exposure and impulsivity with additional
consideration of potential confounders including maternal nonverbal reasoning abilities (Boucher et al..
2012a).
3.5.2.2.2 Toxicological Studies of Impulsivity
The associations described between higher BLL and greater impulsivity in children are supported
by findings in animals for Pb-induced increases in perseveration and impaired ability to inhibit
inappropriate responses. In animals, these effects are supported by studies reviewed in the 1986 and 2006
Pb AQCDs (U.S. EPA, 2006, 1986) and studies incorporated into the 2013 Pb ISA. Animal studies
provide more consistent evidence for the effects of Pb exposure on impulsivity than on sustained
attention. As mentioned earlier, behaviors displayed by animals in a variety of tests can be identified as
reflecting impulsivity. These include tests of differential reinforcement of low rates of responding, fixed
interval (FI) schedule performance, FI with extinction, or fixed ratio (FR)/waiting-for-reward. Greater
impulsivity is indicated by premature responses, decreased pause time between two scheduled events, and
increased perseveration.
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Behaviors observed in tests of operant conditioning with FI reinforcement schedules have also
been used to indicate impaired learning in animals (Section 3.5.1.3.2), and the interactions observed
between Pb exposure and maternal or offspring stress may also apply to effects on impulsivity. Maternal
exposure to 150 ppm Pb with and without stress co-exposure was found to increase overall FI rate and
decrease Post-reinforcement Pause (PRP) in rats. Lifetime (from gestation) Pb exposure resulting in BLLs
of 11-16 (ig/dL increased the overall FI rate without stress co-exposure and decreased PRP with stress co-
exposure (Rossi-George et al.. 2011). suggesting that stress may interact with Pb exposure to affect
attention. Discrimination reversal learning has been shown to be affected by Pb exposure. In these tasks,
an animal is trained to choose between two alternative responses and is then required to reverse the
association. Perseveration or lack of inhibition of the original response can be interpreted to involve
impulsivity. Spatial and non-spatial discrimination reversal was significantly affected in monkeys after Pb
exposure during infancy, after infancy, or continuously from birth, and was exacerbated with distracting
stimuli (Rice. 1990; Rice and Gilbert. 1990b; Gilbert and Rice. 1987). These monkeys had BLLs in the
range of 15-36 (ig/dL, which includes values relevant to this ISA. Hilson and Strupp (1997) found Pb
exposure (Pb acetate in drinking water at GD 1-PND 28, yielding a BLL of 26 (ig/dL in the lower dose
group) in rats slowed reversal learning in an olfactory discrimination task. However, analysis of the
response patterns showed that Pb exposure shortened the perseverative responding phase of reversal
learning and lengthened the post-perseverative phase of chance responding, indicating impairments in
associative ability, not response inhibition. Thus, it is more likely that Pb negatively affected associative
learning rather than impulsivity in this study (Hilson and Strupp. 1997). which is inconsistent with the FI
data in monkeys. Due to the small number of studies following relevant Pb exposures in rodents, it
remains unclear whether Pb affects impulsivity in FI.
The effects of Pb exposure on impulsivity also have been demonstrated in a study reporting that
Pb-exposed animals wait a shorter period of time for reward in FR/waiting for reward testing. In this test,
animals can obtain food by pressing a lever a fixed number of times (FR component). Free food is then
delivered at increasingly longer time intervals, so long as the animal inhibits additional lever presses.
Animals can reset the schedule to return to the FR component at any time. Brockel and Corv-Slechta
(1998) exposed male Long-Evans rats to 0, 50, or 150 ppm Pb acetate in drinking water from weaning,
which produced respective BLLs of <5, 11, and 29 (ig/dL after 3 months of exposure. After 40 days of
exposure, the 150 ppm Pb-exposed rats responded more quickly in the FR component and reset the
schedule (thus shortening the waiting period) more often than did the 50 ppm Pb-exposed rats and
controls. In the waiting component, average wait time was significantly lower in both Pb exposure groups
compared with controls. The rats exposed to 150 ppm Pb also had higher response rates and earned more
reinforcers per session but had a higher response to reinforcement-ratio than did the 50 ppm Pb group and
controls, which indicated less efficient responses.
Weston et al. (2014) used the delayed discounting paradigm following developmental exposure
with or without prenatal restraint stress. The delayed discounting protocol offered animals the choice
between a large reward after a long delay or a small reward after a short delay. Pb increased long-delay
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responding, slowed acquisition of delayed discounting performance, and increased failures almost
exclusively in males. Consistent with (Hilson and Strupp. 1997). these results more likely represent
impaired learning or cognitive flexibility rather than simply increased impulsivity.
In summary, several studies in animals indicate that Pb exposure of rodents and nonhuman
primates from birth or after weaning changes behavior in ways consistent with increased impulsivity,
primarily as indicated by impaired response inhibition. It is also important to note that many of the
measures of impulsivity discussed in this section are sensitive to disruption by impairments in learning
and executive function, which is consistent with several of the studies summarized in Sections 3.5.1.3.2
and 3.5.1.4.2. Some observations of Pb-induced impulsivity in animals were made with BLLs considered
relevant for this ISA. The observations for Pb-induced increases in impulsivity in animals provide support
for associations found in children of higher blood and tooth Pb levels with lower response inhibition and
higher ratings of impulsivity.
Transgenerational Effects of Pb on Impulsivity
The paradigm of combined Pb and stress exposure experienced by a laboratory animal has been
examined with a focus on the common pathway of altered HPA axis and brain neurotransmitter levels.
Studies investigating the interactions between Pb and prenatal stress on learning and memory are
reviewed in Section 3.5.1.3.2. These findings were expanded on in a recent study that investigated the
transgenerational effects of combined Pb and prenatal stress in mice (Sobolewski et al.. 2020). The
authors reported that Pb exposure in the gestating female (F0 generation) resulted in sex-specific effects
in the third filial (F3) generation (no direct exposure to Pb), with F3 females displaying significantly
elevated response rates in an FI schedule of reward compared with control lineages, suggesting an
impulsive behavioral phenotype (Sobolewski et al.. 2020). This and other transgenerational effects were
accompanied by Pb-induced alterations in neurotransmitters, BDNF expression, and DNA methylation.
The authors postulated that lineage effects may be mediated through some combination of maternal
responses to pregnancy, maternal behavior, or epigenetic modifications (Sobolewski et al.. 2020). While
these findings were limited to a single study, they support the possibility that exposure to Pb may
influence the behavior of subsequent generations.
3.5.2.3 Hyperactivity in Children
3.5.2.3.1 Epidemiologic Studies of Hyperactivity in Children
Studies reviewed in the 2006 Pb AQCD (U.S. EPA, 2006) indicated associations between higher
concurrent BLLs or tooth Pb levels and higher parent or teacher ratings of hyperactivity in children aged
6-11 years in the U.S., Asia, and New Zealand (Rabinowitz et al., 1992; Silvaet al., 1988; Gittleman and
Eskenazi, 1983; Needleman et al.. 1979; David et al.. 1976). The case-control or cross-sectional design of
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studies limited understanding of the temporal sequence between Pb exposure and hyperactivity. A
prospective study (Chandramouli et al.. 2009) included in the 2013 Pb ISA also found associations
between BLL and hyperactivity as rated by teachers and parents. Overall, studies indicated associations in
children 3-12 years old with mean concurrent BLLs of 3.7-12 (ig/dL. Studies of recent hyperactivity
symptoms as rated by parents and teachers are discussed in Section 3.5.2.4.
3.5.2.3.2 Toxicological Studies of Hyperactivity
The 2006 Pb AQCD (U.S. EPA, 2006) reviewed the evidence that developmental exposure to Pb
could affect locomotor activity in laboratory animals. Findings summarized in this document included
four studies showing increased activity with developmental Pb exposure and three studies showing no
change in activity. The 2013 Pb ISA (U.S. EPA, 2013) only described one new study in this category,
which showed a decrease in activity in mice after maternal Pb exposure (Leasure et al., 2008). Effects of
developmental Pb exposure on rodent locomotor activity are commonly assessed using an open-field test.
The activities examined vary across studies (e.g., distance traveled, counts of square crossings). Because
there are myriad potential explanations for changes in rodent activity, it can be difficult to draw
conclusions from "simple" tests like open-field. The results from such tests are best interpreted alongside
additional behavioral assays, which, together, may better model the complexity of human behavior.
Conclusions for an effect of developmental Pb exposure on locomotor activity were not reached in earlier
United States Environmental Protection Agency (U.S. EPA) Pb reviews due to mixed results in these
tests. Studies described in the 2013 Pb ISA (U.S. EPA, 2013) and 2006 Pb AQCD (U.S. EPA, 2006) as
observing the effects of maternal Pb exposure, which resulted in mean BLLs no higher than 30 (ig/dL, are
Munoz et al. (1989), Rodrigucs et al. (1996), Moreira et al. (2001), Trombini et al. (2001), De Marco et
al. (2005), and Leasure et al. (2008).
Recent studies (see Table 3-1 IT) observed the activity of early postnatal rodents after
developmental exposures to Pb with varying durations. Tartaglionc et al. (2020) exposed rats to Pb from
pregestation to offspring weaning and tested offspring. They observed a decrease in neonatal spontaneous
activity on PND 10 and no change in spontaneous activity on PND 4, 7, and 12. Another group reported
that two groups of CD 1 mice exposed to either a high or low dose of Pb through lactation exhibited
hyperactivity (PND 7, 11, 15, 19) compared with controls in open-field testing Duan et al. (2017).
Interestingly, when locomotor data from early postnatal studies are pooled, nine out of nine sets of
lactationally exposed animals (BLLs: 9.6-28.9 (ig/dL) tested between PND 14 and 23 were hyperactive
(Duan et al., 2017; De Marco et al., 2005; Moreira et al., 2001; Rodrigucs et al., 1996). These sets
consisted of both male and female rodents, except for one set of only males in Moreira et al. (2001).
Evidence inventory (Section 3.7) also includes recent studies that monitored the effects of
developmental exposure to Pb on immature rodents postweaning. Basha and Reddy (2015) found
decreased locomotor activity in male rats with gestational exposure to Pb when tested on both PND 21
and PND 28. Betharia and Maher (2012) studied open-field behavior in Sprague Dawley rats with
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gestational and lactational exposure to Pb. There were no differences in total square crossings for both
males and females tested at both PND 24 and PND 59. Flores-Montoya and Sobin (2015) saw no effects
on open-field tasks (PND 28) in two groups each of male and female C57BL/6 mice after postnatal
exposure (PND 0-28) to low levels of Pb acetate. Developmentally Pb-exposed rats from the Tartaglione
et al. (2020) study described in the previous paragraph exhibited no change compared with control in
open-field activity when tested on PND 30. Neuwirth et al. (2019a) observed no effect on locomotor
activity in open-field tests (PND 36-45) for two groups of Long-Evans rats with different levels of in
utero and lactational exposure to Pb. Zou et al. (2015) reported that exposure to Pb acetate in drinking
water for 3 weeks (PND 37-58) increased spontaneous locomotor activity in juvenile ICR mice tested on
PND 58. These recent studies do not indicate effects on the activity of rodents when tested in adolescence
after Pb exposure from mothers, and they do not support the findings of hyperactivity in the similarly
exposed and tested mice described by Trombini et al. (2001).
Faulk et al. (2014), Basha et al. (2014), and Wang et al. (2016) evaluated activity in adult rodents
after developmental Pb exposure. Faulk et al. (2014) measured activity and horizontal movements along
with ambulatory activity by adult offspring of dams (Agouti mouse) exposed to Pb for weeks from
pregestation until weaning. Mean BLLs for offspring were not reported; however, maternal BLLs tested
at weaning were below the limit of detection in the control group and 4.1, 25.1, and 32.1 (ig/dL in the
three respective exposure groups of 2.1, 16, and 32 ppm. Overall horizontal activity was different across
Pb exposures in females but not in males. At 9 months, female offspring exposed to 2.1 ppm Pb had
higher average horizontal activity compared with controls. There was a sex-specific difference in
ambulatory measurement (subset of total horizontal activity), with only exposed females showing
significant differences from controls. Ambulatory activity was lower in females at the 32-ppm exposure
level at 3 months versus control offspring. Males did not exhibit significant differences at any time point
or exposure level. Although there was suggestive evidence of differences in the life-course patterns of
vertical activity by exposure among females, neither sex showed statistically significant differences
between Pb-exposed and control offspring. Testing adult rats at 4, 12, and 18 months of age, Basha et al.
(2014) found consistent decreases in locomotor activity associated with lactational Pb exposure. While
mean BLLs at these testing periods were shown to be below the 30 (ig/dL limit for PECOS relevance, it is
notable that the mean BLL for this group of animals was determined to be 49.5 (ig/dL at PND 45. Wang
et al. (2016) measured distance traveled in the open field by Sprague Dawley rats aged 116-122 days
after adolescent Pb exposure in drinking water from PND 24 to 56. They observed no effect of this
exposure.
Evidence of Pb exposure-induced intergenerational effects on rodent behavior was also reviewed
in the 2006 Pb AQCD (U.S. EPA, 2006). Trombini et al. (2001) observed increased open-field
ambulation in F2 generation rats derived from female offspring of Pb-treated pregnant mothers. Recently,
Sobolewski et al. (2020) exposed F0 mice to Pb during pregnancy and lactation, bred offspring with
unexposed mates for two generations (F1 and F2), and then evaluated behavior in the F3 generation. F3
females demonstrated a small increase in locomotor activity, regardless of lineage.
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Overall, there are still mixed indications on locomotor activity from Pb exposure studies with
BLLs <30 |ig/dL. which may be due to differential dosing, timing of exposures, and activity
measurements. However, in a rare set of four individual studies wherein these critical factors were
analogous, Pb exposure during lactation induced hyperactivity in rodents when tested within a PND 14 to
23 window (Duan et al.. 2017; De Marco et al.. 2005; Moreiraet al.. 2001; Rodrigues et al.. 1996). Pb-
induced hyperactivity in rodents provides some support for hyperactivity observed in children but may be
more appropriately interpreted in the context of additional behavioral assays.
3.5.2.4 Parent and Teacher Ratings of ADHD-related Behavior
In addition to finding associations with attention, impulsivity, and hyperactivity, epidemiologic
studies also found associations between higher concurrent BLLs and higher parent and teacher ratings of
ADHD-related behaviors, calculated as a composite of the various behaviors evaluated in the diagnosis of
ADHD (see Section 4.3.3.1 of the 2013 Pb ISA (U.S. EPA. 2013)). Most of these studies were limited
due to their cross-sectional design and lack of validation of ADHD ratings with clinical diagnosis.
Although diagnostic guidelines for ADHD exist, the exact criteria or specific behaviors required can vary.
Thus, within studies, there were variations among subjects in the types of behaviors they displayed that
led to a diagnosis of ADHD. Further the available studies considered age, sex, and SES or parental
education but generally not both as potential confounders, and none of the studies considered parental
caregiving quality.
Recent longitudinal studies add to the body of evidence examining the association between
prenatal and childhood Pb exposure and parent/teacher-rated ADHD symptoms in populations with
relatively low blood Pb concentrations (<5 (ig/dL). This group of studies includes some that found
associations with hyperactivity using the SDQ, which is a screening questionnaire that includes five
domains. Sioen et al. (2013) analyzed data from the Flemish Environment and Health Study (FLEHS I,
2002-2006), a birth cohort comprising mother-infant pairs to examine the association between cord blood
Pb and ADHD-related behaviors for 281 infants whose parents returned the SDQ (26.4%). A positive
association of cord blood Pb concentration with hyperactivity score >7 was observed (OR: 2.94 [95% CI:
1.17, 7.38 per log ug/dL increase in BLL]). In another study using this assessment instrument, Fruh et al.
(2019) analyzed data from mother-child pairs participating in Project Viva, a longitudinal birth cohort in
eastern Massachusetts. Maternal blood Pb concentration in erythrocytes was measured during the second
trimester of pregnancy and parents rated their child's behavior using the SDQ in mid-childhood (median
7.7 years). The associations (i.e., |3 coefficients) with the parent and teacher-rated hyperactivity
component of the SDQ were 0.10 (95% CI: -0.21, 0.41) and 0.20 (95% CI: -0.24, 0.64), respectively.
While behavior assessments and maternal blood Pb measurements were available for fewer than half of
Project Viva participants, important confounders including HOME score, maternal IQ, and parental
education were considered in this study.
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Several of these longitudinal epidemiologic studies used the Behavior Assessment System for
Children (BASC) to assess both the behaviors and emotions of children (Reynolds and Kamphaus. 2015).
BASC-2 includes individual subscales for attention and hyperactivity as well as an overall behavioral
skills index (BSI) composite score. Specific rating scales and forms related to attention, hyperactivity, and
impulsivity are emphasized in this section (e.g., clinical scales such as "attention problems" or
"hyperactivity" on the teacher or parent rating scale forms).
Horton et al. (2018) analyzed data from the Early Life Exposure in Mexico to Environmental
Toxicants (ELEMENT) Project birth cohort in Mexico City to determine the association of weekly tooth
Pb concentration (prenatal through 1 year postnatal) with BASC-2 scores assessed between 8 and 11
years old. Distributed lag models were used to identify specific time windows of increased risk due to Pb
exposure. Tooth Pb concentration estimated to correspond with the 8 to 11 months postnatal period was
associated with parent-rated behavioral symptoms overall (|3 = 0.22 units [95% CI: 0.06, 0.38] per natural
log unit increase in dentine Pb concentration), and hyperactivity (|3 = 0.19 units [95% CI = 0.02, 0.37] per
natural log unit increase in dentine Pb concentration) after adjustment for gestational age and maternal
education. Approximately 12% of the original cohort was enrolled in this study; participants differed with
respect to several characteristics including child birth weight and maternal IQ. Rasnick et al. (2021)
conducted a study that estimated monthly air Pb exposure. The authors also aimed to identify sensitive
time windows of exposure; however, they attempted to distinguish exposure to Pb in air by controlling for
concurrent BLL (age 12 years) in their analysis of the Cincinnati Study of Allergy and Air Pollution study
data. Air Pb exposure was estimated using validated land use regression models and behavioral outcomes,
including attention and hyperactivity, were assessed using BASC-2 administered at age 12. Distributed
lag models to predict outcome responses based on current and past (i.e., lagged) predicted air Pb
exposures did not identify associations during any of the lifestages examined. Models were adjusted for
community deprivation, residential greenspace, and elemental carbon attributable to traffic (ECAT), in
addition to concurrent BLL.
In addition to examining attention using Couriers' (Section 3.5.2.1.1), Ruebner et al. (2019)
evaluated the association of BLLs with parent-rated attention and hyperactivity symptoms on BASC-2.
This study was unique in that it enrolled children with CKD. Associations with parent ratings did not
persist in models that controlled for potential confounders including race, poverty, maternal education,
and clinical factors related to CKD. The median BLL in this study was 1.2 (ig/dL.
Several other instruments, including FBB-ADHS, the Child Behavior Checklist (CBCL), the
Disruptive Behavior Disorder (DBD) rating scale, the Barkley Adult ADHD-IV Rating Scale (BAARS),
Couriers' Rating Scale (CRS), the Strengths and Weaknesses of ADHD Symptoms and Normal Behavior
Scale (SWAN), and DuPaul's ADHD rating scale were used to assess total ADHD in recent prospective
or case-control studies. These rating scales are generally reliable and valid instruments that predict
functionally important outcomes (Desrochers-Couture et al.. 2019; Fruh et al.. 2019; Nigg et al.. 2016;
Hong et al.. 2015; Gittleman and Eskenazi. 1983).
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Neugebauer et al. (2015) conducted an analysis of the Duisburg birth cohort data to determine the
association of maternal BLL at 32 weeks gestation with parent-rated ADHD behaviors in childhood
(average age 9.5 years old) assessed using FBB-ADHS. Maternal blood Pb was associated with overall
ADHD symptoms (|3 = 1.06 [95% CI: 1.01, 1.12]), with the strongest association observed forthe
impulsivity component (|3 = 1.13 [95% CI: 1.06, 1.22]). These associations were observed after
adjustment for confounders including parental education, but not SES.
Liu et al. (2014b) examined the association of early childhood blood Pb concentration at 3, 4, or 5
years old (mean: 6.8 (ig/dL) with parent and teacher ratings of ADHD behaviors among Chinese school
children at age 6 using CBCL and the Caregiver-Teacher Report Form (C-TRF). The outcome was
modeled as a continuous and also as a dichotomous variable (i.e., clinically significant behavior problems
when T-score >60). The associations (i.e., |3) between increased blood Pb concentrations (per (ig/dL) and
ADHD behavior problems were 0.001 (95% CI: -0.002, 0.002) for problems reported on CBCL and 0.07
(-0.18 to 0.32) for behavior problems reported on C-TRF. The associations (i.e., OR) with clinically
significant ADHD behavior reported by parents on CBCL were 1.08 (95% CI: 0.99, 1.18) among children
overall, 1.04 (95% CI: 0.94, 1.16) among boys, and 1.15 (95% CI: 0.98, 1.35) among girls. The
participation rate was 81% in this study. Models were adjusted for confounders including parental
caregiving quality but not SES. Another prospective study evaluated the association of Pb exposure with
caregiver ratings on CBCL. In this study of adolescents, Winter and Sampson (2017) examined the
relationship between average BLLs in childhood (6 years old or younger) with impulsivity between 16 to
18 years old. These authors found a 0.06 SD (95% CI: 0.01, 0.12) increase in impulsivity score, after
adjustment for caregiver education and SES. Participants were originally enrolled in the mid-1990s and a
random sample of those that continued to participate in 1999 and 2002 was randomly selected for this
study, with 67% of those selected agreeing to participate.
Choi et al. (2016) investigated the association of childhood BLLs (geometric mean BLL =1.56
(ig/dL) with parent-rated ADHD symptoms later in childhood assessed using DuPaul's ADHD rating
scale. Approximately 72% (n = 2,159) of 2,967 eligible participants provided blood Pb measurements and
ADHD assessments, and 2052 were free of ADHD symptoms at baseline. A positive association between
childhood blood Pb and the development of ADHD symptoms at the 2-year follow-up visit was observed
in this study (RR: 1.55 [95% CI: 1.00, 2.40] >2.17 versus <2.17 (ig/dL) after adjustment for residential
area, household income, parental marital status, family history of psychiatric disorders, preterm birth, and
birth weight. A stronger association was observed among children with higher BLLs and who resided in a
single parent home (RR: 3.57 [95% CI 1.60, 7.98]).
Another longitudinal analysis examined the association of both cord BLL (mean: 4.7 (ig/dL) and
childhood (mean 2.7 (ig/dL) BLL with ADHD symptoms among Inuit children in Quebec (Boucher et al..
2012b). In this study, teachers completed the DBD rating scale to indicate Diagnostic and Statistical
Manual of Mental Disorders (DSM)-IV symptoms of ADHD inattentive type and ADHD hyperactive-
impulsive type. Associations between child concurrent log-transformed BLL and hyperactive/impulsive-
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type ADHD symptoms assessed using DBD were observed (OR = 4.01 [95% CI: 1.06, 5.23] tertile 2
versus tertile 1; OR = 5.52 [95% CI: 1.38, 22.12] tertile 3 versus tertile 1). Inattentive-type ADHD
symptoms on the DBD were not associated with Pb exposure. (Desrochers-Couture et al., 2019) extended
this study by conducting a mediation analysis to estimate the direct and indirect associations of childhood
BLLs with adolescent externalizing behaviors, including ADHD symptoms assessed using BAARS. The
study found an association between childhood BLL and child hyperactivity and impulsivity, assessed by
teachers on CBCL (|3 = 0.45 [95% CI: 0.13, 0.78]). Neither a direct (|3 = 0.09 [95% CI: -0.11, 0.28]) nor
an indirect (|3 = -0.02 [95% CI: -0.06, 0.03]) association with adolescent ADHD symptomology assessed
using BAARS was observed. A wide array of covariates was considered as potential confounders
including both maternal education and SES.
Hong et al. (2015) found an association between blood Pb concentration and higher parent and
teacher-rated ADHD-RS symptoms (|3 = 1.04 [95% CI: 0.18, 1.90] and |3 = 1.90 [95% CI: 0.74, 3.05],
respectively) in a cross-sectional analysis of Korean school children from 8 to 11 years old after
adjustment for demographic factors (age, sex, residential region, paternal education level, and SES). This
association remained positive but was attenuated in models additionally adjusted for FSIQ, Mn, and Hg (|3
= 0.68 [95% CI: -0.20, 1.56] and |3 = 1.49 [95% CI: 0.32, 2.67], parent- and teacher-rated symptoms,
respectively). The mean BLL in this study was 1.80 (ig/dL. Associations indicating an increase in
commission errors on CPT were also observed.
Nigg et al. (2016) conducted a case-control study of children from Michigan (mean BLL = 0.74
(ig/dL (cases) and 0.94 (ig/dL controls). In this study, ADHD composite indices were derived for (1)
inattention/disorganization and (2) composite hyperactivity-impulsivity using relevant scales of the
DuPaul, Couriers', and SWAN scales. This study found an interaction between the hemochromatosis gene
(HFE) C282Y genotype, which is involved in iron metabolism, and BLL in predicting parent and teacher
reports of hyperactivity-impulsivity but not inattention. For example, the association between z scores of
BLL and hyperactivity was significantly stronger among those with the HFE C282Y mutation (|3 = 0.74,
[95% CI: 0.52, 0.96]) compared with those with the wild type genotype (|3 = 0.28 [95% CI: 0.15, 0.41]).
This study also found an interaction between z scores of logio-transformed BLL and sex (association
larger in boys) in predicting parent and teacher-rated hyperactivity and impulsivity but not attention.
3.5.2.4.1 Summary
Cross-sectional studies in the 2013 Pb ISA found associations between higher concurrent BLL
and higher parent and teacher ratings of ADHD-related behaviors, calculated as a composite of the
various behaviors that are evaluated in the diagnosis of ADHD (U.S. EPA, 2013). The evidence from
prospective studies was limited to Chandramouli et al. (2009), which found associations between BLL
and hyperactivity as rated by teachers and parents. Parent and teacher ratings generally considered SES or
parental education, but typically not both, as potential confounders. None of the studies considered
parental caregiving quality. Recent longitudinal studies that established the temporality between the
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exposure and the outcome add to the body of evidence examining the association between prenatal and
childhood Pb exposure and parent/teacher-rated ADHD symptoms in populations with relatively low
blood Pb concentrations (<6 (.ig/dL). Across studies, associations were observed with tooth Pb
concentration that were measured in dentin and generally reflect early childhood Pb exposure, childhood
BLLs, and maternal or cord (2-5 (ig/dL) BLLs. Studies of caregiver-reported ADHD symptoms generally
report associations with composite indices (Choi et al., 2016; Hong et al., 2015; Neugebauer et al., 2015;
Liu et al., 2014b; U.S. EPA, 2013), and there is some evidence indicating that the associations with
impulsivity and hyperactivity symptoms (Desrochers-Couture et al., 2019; Fruh et al., 2019; Horton et al.,
2018; Winter and Sampson, 2017; Nigg et al„ 2016; Neugebauer et al., 2015; Sioen et al„ 2013; Boucher
et al., 2012b) are stronger than the associations with inattention symptoms. The majority of recent studies
were prospective and generally reported moderate or high participation rates. Some studies addressed the
validity of caregiver assessed outcomes by evaluating internal consistency (Rasnick et al., 2021;
Desrochers-Couture et al., 2019), and Nigg et al. (2016) addressed reliability and validity concerns by
using structural equation modeling to create latent factors for inattention and hyperactivity-impulsivity for
each informant. Rating scales used in these studies are generally reliable and valid instruments that
predict functionally important outcomes (Desrochers-Couture et al., 2019; Fruh et al., 2019; Nigg et al.,
2016; Hong et al„ 2015; Gittleman and Eskenazi, 1983). Confounder adjustment remains somewhat
inconsistent across studies, although Liu et al. (2014b) and Fruh et al. (2019) adjusted for the quality of
parental caregiving, Choi et al. (2016) adjusted for family history of psychiatric disorders, and several
considered both SES and parental education (Desrochers-Couture et al., 2019; Ruebner et al., 2019;
Horton et al„ 2018; Winter and Sampson, 2017; Boucher et al., 2012b). There is uncertainty regarding the
patterns of exposure that are associated with BLLs in older children because they may be influenced by
higher past exposure.
3.5.2.5 Clinically Diagnosed ADHD
In the 2013 Pb ISA, results from a small body of cross-sectional studies indicated associations
between concurrent BLL and the prevalence of ADHD symptom ratings (Section 3.5.2.4) and clinically
diagnosed ADHD in children aged 4-17 years. The temporal relationship between Pb exposure and
ADHD was not established in these studies, and concurrent blood Pb concentrations in older children may
reflect higher past exposures. Additionally, some ADHD symptom rating studies lacked outcome
validation, and confounding was inconsistently addressed across studies. Therefore, the evidence
specifically for these total ADHD index ratings and clinically diagnosed ADHD were emphasized less in
the 2013 Pb ISA (U.S. EPA, 2013) than evidence for individual behaviors when drawing conclusions
about the effects of Pb exposure on attention, impulsivity, and hyperactivity.
Recent studies add to the evidence and address some of the uncertainties pertaining to the studies
included in the 2013 Pb ISA. Notably, Ji et al. (2018) analyzed data from the Boston Birth Cohort (1479
mother-infant pairs) to examine the association of early childhood Pb exposure (i.e., earliest (< age 4)
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blood Pb concentration recorded during routine screening) with the development of ADHD later in
childhood. ADHD was assessed using electronic medical records (International Classification of Diseases
[ICD]-9 codes: 314.0, 314.00, 314.01, 314.1, 314.2,314.8, and 314.9, or ICD-10 codes: F90.0, F90.1,
F90.2, F90.8, and F90.9). Several important potential confounders (i.e., parental education, SES but not
quality of parental caregiving) were controlled for in the analysis, and child sex, maternal high-density
lipoprotein (HDL), and maternal stress were considered as potential effect modifiers. Ji et al. (2018)
analyzed the association modeling blood Pb concentration as continuous and as categorical variables.
When blood Pb was analyzed using three categories, the OR comparing children with BLLs between 2
and 4 (ig/dL to children with BLLs <2 (ig/dL was 1.08 (95% CI: 0.81-1.44). The OR comparing children
with BLLs between 5 and 10 (ig/dL to children with BLLs <2 (ig/dL was 1.73 (95% CI: 1.09-2.73).
When blood Pb was modeled as a continuous variable, the OR was 1.12 [95% CI: 1.00, 1.25) per ug/dL
increase in BLL. Sex-stratified analyses comparing children with BLLs between 5 and 10 (ig/dL to
children with BLLs <5 (ig/dL indicated no association among girls (OR = 0.68 [95% CI: 0.27, 1.69]) and
a strong association among boys (OR = 2.49 [95% CI: 1.46-4.26]). Joint analyses indicated a 10-fold
increase in the magnitude of the association between childhood blood Pb concentration and ADHD
diagnosis among those with multiple risk factors (i.e., male sex, inadequate maternal HDL, higher
maternal stress).
Several additional recent studies also extend the evidence. Park et al. (2016) conducted a hospital
based case-control study in Busan, South Korea comparing the odds of higher blood Pb concentration
among diagnosed ADHD cases, which were confirmed using the Korean version of the Kiddie Schedule
for Affective Disorders and Schizophrenia Present and Lifetime (K-SADS-PL-K), to the odds of higher
blood concentration among controls that were frequency matched by age and sex and adjusted for other
potential confounders. Blood Pb was measured when the cases and controls were recruited into the study.
Higher blood Pb concentration was associated with increased risk of ADHD (OR: 1.60 [95 % CI: 1.04-
2.45] per unit increase in log BLL); however, blood Pb concentrations were not associated with ADHD-
RS score or CPT profiles among the ADHD cases. In a smaller case-control study that examined the
association of childhood Pb exposure with diagnosed ADHD among children living near a former smelter
in Omaha, Nebraska, Kim et al. (2013a) found a positive association (OR: 2.52 [95% CI: 1.07, 5.92] per
unit increase in natural log BLL).
In a recent cross-sectional analysis of NHANES (2003-2004) data, Geier et al. (2018) examined
the association of concurrent blood Pb concentration with self-reported doctor diagnosed attention deficit
disorder (ADD) among children and adolescents 10-19 years old. This study observed a positive
association between concurrent blood Pb concentration and ADD after adjusting for age, race, sex, and
SES (OR: 1.29 [95% CI: 1.03, 1.55]). In a previous analysis Braun et al. (2006) found an association in
children aged 4-15 years participating in NHANES (1999-2002). ADHD ascertained by the parent report
of ADHD diagnosis is subject to reporting bias; however, the examination of multiple risk factors and
outcomes in NHANES reduces the likelihood of biased participation and reporting of ADHD by parents
of children specifically with higher Pb exposure.
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3.5.2.5.1 Summary
The 2013 Pb ISA assessed a small body of cross-sectional studies that examined the associations
between concurrent BLLs and the prevalence of clinically diagnosed ADHD. The temporal relationship
between Pb exposure and ADHD was not established in these studies and it was noted that concurrent
blood Pb concentration in older children potentially reflects higher past exposures. As noted in the 2013
ISA, clinically diagnosed ADHD was emphasized less than evidence for individual behaviors in drawing
conclusions about the effects of Pb exposure on attention, impulsivity, and hyperactivity in the 2013 Pb
ISA (U.S. EPA, 2013). Further, the available studies did not consistently adjust for SES, parental
education, and quality of parental caregiving. A small number of recent studies add to the evidence
showing consistent associations between Pb exposure and diagnosed ADHD. One recent epidemiologic
study (Ji et al., 2018) addressed several of the uncertainties identified in the literature included in the 2013
Pb ISA. Specifically, this study employed a prospective design, and adjusted for parental education and
SES (although not quality of parental caregiving). Notably, ADHD was ascertained using ICD codes
recorded on electronic records and ADHD type was not distinguished in this study.
3.5.2.6 Relevant Issues for Interpreting the Evidence Base
3.5.2.6.1 Lifestages
Environmental exposures during critical lifestages spanning from childhood into adolescence can
affect key physiological systems that orchestrate brain development and plasticity (see Section 3.4.1.6.4
of U.S. EPA (2013)). Epidemiologic studies examined in the 2013 Pb ISA consistently showed that BLLs
measured during various lifestages and time periods, including the prenatal period, early childhood, later
childhood, and averaged over multiple years, are associated with attention decrements, impulsivity, and
hyperactivity in children. These observations of Pb-associated elevated risk are well supported by
findings in animals that prenatal and early postnatal or lifetime Pb exposures alter brain development via
changes in synaptic architecture (Section 4.3.10.4 of U.S. EPA (2013)) and neuronal outgrowth (Section
4.3.10.10 of U.S. EPA (2013)), potentially leading to increases in impulsivity (Section 4.3.3.1 of U.S.
EPA (2013)). Potential mechanisms of lifestage-specific sensitivities are further reviewed in Section 3.3.
Recent studies support this conclusion from the 2013 Pb ISA.
A limited number of epidemiologic studies employed methods designed to further elucidate
critical lifestages for Pb exposure but did not change the overall conclusion in the 2013 Pb ISA. Horton et
al. (2018) used distributed lag models to identify specific time windows of increased Pb-associated
externalizing behaviors. This study found that tooth Pb concentration corresponding with the 8 to 11
months postnatal period was associated with parent-rated behavioral symptoms overall (0.22 units [95%
CI: 0.06, 0.38] per natural log unit increase in dentine Pb concentration), and hyperactivity (|3 = 0.19 units
[95% CI = 0.02, 0.37] per natural log unit increase in dentine Pb concentration) after adjustment for
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gestational age and maternal education. In another study, Rasnick et al. (2021) also aimed to distinguish
critical windows of exposure to Pb, focusing on Pb concentration in air by controlling for concurrent BLL
(age 12 years). Air Pb exposure was estimated using validated land use regression models and behavioral
outcomes, including attention and hyperactivity, were assessed using BASC-2 administered at age 12.
Distributed lag models to predict outcome responses based on current and past (i.e., lagged) predicted air
Pb exposures did not identify associations during any of the lifestages examined.
3.5.2.6.2 Public Health Significance
The strongest evidence indicating a causal relationship between Pb exposure and attention,
impulsivity, and hyperactivity assessed in the 2013 Pb ISA was derived from studies that relied on
neuropsychological testing (U.S. EPA, 2013). Domain-specific neuropsychological assessments of
attention, impulsivity, and hyperactivity have strong psychometric properties and rigorous validation;
however, deficits on these neuropsychological tests do not directly correspond to a diagnosis of ADHD
nor do they necessarily predict long-term consequences that might be associated with some types of
ADHD. Studies that evaluated the association of Pb exposure with behavioral symptoms of ADHD
assessed using teacher and parent ratings contributed to the overall evidence in the 2013 Pb ISA, but the
limitations of these studies were noted. The bulk of the recent evidence comprises prospective studies of
parent or teacher ratings of ADHD behavioral symptoms. The recent studies addressed some uncertainties
in the previous ISA related to the temporal association of the exposure with the outcome and controlled
for potential confounding. Studies of diagnosed ADHD are also subject to limitations. Although
diagnostic guidelines for ADHD exist, the exact criteria or specific behaviors required for diagnosis may
vary across studies. The recent study by Ji et al. (2018) addressed several of the uncertainties regarding
the association of Pb exposure with clinical ADHD. This study was prospective in design, assessed early
childhood BLL (<4 years old), and adjusted for parental education and SES (although not quality of
parental caregiving); however, ADHD was ascertained using ICD codes recorded on electronic records
and ADHD type was not distinguished.
3.5.2.6.3 Potentially At-Risk Populations
Sex
Studies examining sex as an at-risk factor for attention, hyperactivity and impulsivity outcomes
were not assessed in the 2013 Pb ISA. A recent study by Nigg et al. (2016) found an interaction between
BLL and sex in predicting parent and teacher-rated hyperactivity and impulsivity but not attention. The
association was larger in boys in this study.
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Maternal Smoking
Maternal smoking during pregnancy was examined in a study of children's concurrent BLLs and
the prevalence of ADHD among children aged 8-15 years. An interaction was observed between
children's current BLLs and prenatal tobacco smoke exposure; those children with high Pb levels and
prenatal tobacco smoke exposure had the highest odds of ADHD (Frochlich et al.. 2009). Recent studies
have not examined maternal smoking as an at-risk factor.
Co-exposure to Other Metals or Chemicals
Studies examining other metals as an at-risk factor for attention, hyperactivity and impulsivity
outcomes were not assessed in the 2013 Pb ISA. Some recent studies adjusted for other metals or
chemicals (e.g., PCBs) (Ethicr et al.. 2015; Tatsuta et al.. 2014). and effect modifications were observed
in other studies (Yorifuji et al.. 2011). For example, Yorifuji et al. (2011) found a less-than-additive
interaction between cord Pb and Hg concentrations. Specifically, a lower digit span forward score on the
WISC-R (|3 = -1.70 [95% CI: -3.12, -0.28] per log-transformed BLL) at age 7 and a lower digit span
backward score on the WISC-R (|3 = -2.73 [95% CI: -4.32, -1.14] per log-transformed BLL) at age 14
were observed among children with the lowest Hg exposure.
Gene-Environment Interactions
Studies examining gene-environment interactions in the context of attention, hyperactivity and
impulsivity outcomes were not assessed in the 2013 Pb ISA. Interactions between child BLL and genes
that regulate neurodevelopmental processes were observed in studies of attention (Choi et al.. 2020;
Roonev et al.. 2018). Genes that were implicated included variants of GRIN2A and GRIN2B and
genotypes involved in the regulation of noradrenergic pathways. In addition, Nigg et al. (2016) found an
interaction between the HFE C282Y genotype and BLL in predicting parent and teacher reports of
hyperactivity-impulsivity but not inattention. Specifically, the association between z scores of BLL and
hyperactivity was significantly stronger among those with the HFE C282Y mutation (|3 = 0.74 [95% CI:
0.52, 0.96]) compared with those with the wild type genotype (|3 = 0.28 [95% CI: 0.15, 0.41]).
3.5.2.7 Summary and Causality Determination: Attention, Impulsivity, and
Hyperactivity
Attention, hyperactivity, and impulsivity are included within the ADHD domain of externalizing
behaviors. Although not studied as extensively as cognitive function, several epidemiologic studies have
examined the relationship between Pb exposure in children and attention, impulsivity, and hyperactivity
in children and young adults.
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The majority of these studies examined attention, and some also examined impulsivity or
hyperactivity. Thus, the focus of the evaluation is on the evidence related to attention, but the evaluation
also draws on coherence with evidence for impulsivity and hyperactivity, including evidence in animals
and that suggesting potential modes of action. The collective epidemiologic evidence base for attention in
children comprises many prospective and cross-sectional studies, which were also reviewed in the 2006
Pb AQCD and the 2013 Pb ISA and some recently published studies. Most of these studies reported
associations between childhood blood Pb or tooth Pb levels, that reflected early postnatal Pb exposure,
and attention decrements, impulsivity, and hyperactivity (Table 3-7E). A small number of recent
longitudinal studies contributed to this evidence. Not all results were uniform with regard to precision and
the magnitude of the association, but results mostly showed a pattern of attention decrements, impulsivity,
and hyperactivity with higher blood or tooth Pb levels.
Whether prospective, cross-sectional, or longitudinal, most studies relied on population-based
recruitment from prenatal clinics, hospitals at birth, or schools and reported moderate to high
participation. Several of the studies reviewed in the previous ISA demonstrated increased loss-to-follow-
up in certain groups (e.g., lower SES or HOME scores), which has the potential to introduce selection
bias and reduce the generalizability of findings. A strong indication that participation in the study was
biased to those with higher BLLs and greater deficits in attention, hyperactivity, or impulsivity was not
observed. Recent studies incorporated adjustments for these and other covariates. Repeated testing in
children was common but the consistent pattern of association observed across the ages, BLL, and
behavioral outcomes examined increases confidence that the evidence is not unduly biased by the
increased probability of finding associations by chance alone. Coherence with animal studies, which are
less vulnerable to confounding, further supports the pattern of associations described in the preceding
sections.
The strongest epidemiologic evidence indicating an association of Pb exposure with inattention
and hyperactivity is described in the 2013 Pb ISA U.S. EPA (2013). Prospective studies showed strong
support for an association between Pb exposure (range: 7-14 (ig/dL) and decreased scores on
neuropsychological tests and parent/teacher ratings of attention and hyperactivity. Cross-sectional studies
from the 2013 Pb ISA generally corroborated these observations. Studies of impulsivity in children in the
2013 Pb ISA were limited by their quantity and lack of temporality but generally indicated associations of
Pb exposure with worse scores on tests of response inhibition and on parent/teacher ratings of impulsivity
in cross-sectional analyses. A small number of recent prospective studies with mean maternal and cord
BLLs <5 (ig/dL report associations with some measures of inattention (Ethier et al„ 2015; Neugebaueret
al., 2015). In addition, recent analyses of Inuit children add support for the relationship between child and
cord BLL and impulsivity (Boucher et al., 2012a).
Most of the aforementioned studies of parent and teacher ratings of ADHD-related behaviors in
the 2013 Pb ISA were largely cross-sectional in design. The evidence from prospective studies was
limited to Chandramouli et al. (2009), which found associations between BLLs and hyperactivity as rated
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by teachers and parents. The available studies considered SES or parental education but generally not
both as potential confounders, and none of the studies considered parental caregiving quality. The bulk of
the recent evidence comprises prospective studies that establish the temporality of the association
between Pb exposure and parent or teacher ratings of ADHD symptoms and clinical ADHD. Across
studies, associations were observed with tooth Pb concentration, childhood (<6 |ig/dL). and maternal or
cord (2-5 (ig/dL) BLLs. Studies of caregiver-reported ADHD symptoms generally reported associations
with composite indices, and there is some evidence that the associations with impulsivity and
hyperactivity symptoms are stronger than the associations with inattention symptoms. The majority of the
recent studies were prospective and generally reported moderate or high participation rates. Some studies
addressed the validity of caregiver assessed outcomes by evaluating internal consistency (Rasnick et al.,
2021; Desrochers-Couture et al„ 2019), and Nigg et al. (2016) addressed reliability/validity concerns by
using structural equation modeling to create latent factors for inattention and hyperactivity-impulsivity for
each informant. Confounder adjustment has become more consistent across recent studies. In addition to
the studies relying on parent and teacher behavior ratings, a small number of recent studies add to the
evidence showing consistent associations between Pb exposure and diagnosed ADHD. One recent
epidemiologic study (Ji et al., 2018) addressed several of the uncertainties identified in the literature
included in the 2013 Pb ISA. Specifically, this study employed a prospective design, assessed early
childhood BLL, and adjusted for parental education and SES (although not quality of parental
caregiving). Notably, in this study, ADHD was ascertained using ICD codes recorded on electronic
records and ADHD type was not distinguished. Uncertainty remains regarding the patterns of exposure
associated with BLLs in older children because they may be influenced by higher past exposures.
The findings from epidemiologic studies are generally coherent with findings of studies in
experimental animals. Available evidence in animals supports the effect of developmental Pb exposure in
rodents and nonhuman primates on behavioral measures consistent with increased impulsivity, primarily
indicated by impaired response inhibition. Measures of impulsivity are additionally sensitive to disruption
by impairments in learning and executive function, which is consistent with several of the studies
summarized in Sections 3.5.1.3.2 and 3.5.1.4.2. While no recent animal toxicological studies of attention
are available, evidence from the 2013 Pb ISA demonstrates Pb-induced decreases in attention in rodents
and monkeys, although results are not entirely consistent across studies. Pb has been observed to have
mixed effects on locomotor activity in rodents, but several studies have demonstrated Pb-induced
hyperactivity in rodents during the postnatal phase, which provides some support for the epidemiologic
findings.
In summary, the total body of evidence evaluated in this and previous assessments is
sufficient to conclude that there is a causal relationship between Pb exposure and attention,
impulsivity, and hyperactivity. This conclusion reflects the consistency of the results from
epidemiologic studies of externalizing behaviors in children and young adults, incorporating various
objective neuropsychological tests and reporting from teachers and parents, which are generally reliable
and valid instruments that predict functionally important outcomes. The conclusion also incorporates the
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coherence of evidence across epidemiologic and toxicological studies of externalizing behaviors and
biological plausibility provided by studies that outline pathways by which Pb may interfere with the
proper development, connectivity, and function of systems underlying externalizing behaviors.
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Table 3-3 Summary of evidence indicating a causal relationship of Pb exposure with attention, impulsivity,
and hyperactivity
Rationale for
Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with Effects0
Consistent
associations from
multiple prospective
epidemiologic
studies with relevant
BLLs
Evidence from prospective
studies for attention decrements
and hyperactivity in association
with prenatal (maternal or cord),
early childhood, and lifetime
blood Pb and tooth Pb levels in
children ages 7-17 yr and young
adults 19-20 yr in the United
States, United Kingdom
Australia, New Zealand.
Evidence from prospective
studies of parent or teacher-
rated ADHD composite
symptom indices derived from
widely used, structured
instruments.
Burns et al. (1999)
Ris et al. (2004)
Ferqusson et al. (1993)
Bellinger et al. (1994a)
Chandramouli et al. (2009)
Leviton et al. (1993)
Section 4.3.3.1, U.S. EPA (2013)
Neuqebauer et al. (2015)
Ethier et al. (2015)
Choi et al. (2016)
Neuqebauer et al. (2015)
Liu et al. (2014b)
Blood Pb:
Means 2 to 8.3 |jg/dL (prenatal maternal or cord),
8.3 |jg/dL (age 6 yr), 13.4 |jg/dL (age 3-60 mo), 14
|jg/dL (lifetime avg to age 11-13 yr)
Group with age 30 mo >10 |jg/dL
Tooth Pb (ages 6-8 yr): Means: 3.3, 6.2 |jg/g
Childhood BLL <6 |jg/dL; maternal and cord BLL 2-5
pg/dL.
Ratings for impulsivity and
hyperactivity more strongly
associated with Pb exposure
Sioen et al. (2013)
Fruh et al. (2019)
Horton et al. (2018)
Neuqebauer et al. (2015)
Winter and Sampson (2017)
Desrochers-Couture et al. (2019)
Boucher et al. (2012b)
Niqq et al. (2016)
Childhood BLL <6 |jg/dL; maternal and cord BLL 2-5
pg/dL.
Prospective analysis found
associations with impulsivity in
Inuit children
Boucher et al. (2012a)
Mean 4.7 (cord), 2.7 (concurrent, average age 11.3
yr)
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Rationale for
Causality
Key Evidence13
References'3
Pb Biomarker Levels Associated with Effects0
Determination3
Prospective analysis find
Ruebner et al. (2019)
Child blood Pb ~2 yr before outcome assessment at
association with attention
age 4-18 yr
decrements in children with CKD
Limited evidence
Co-exposure to Hg modified the
Yorifuii et al. (2011)
evaluates the
risk of Pb-associated effects on
potential modification
attention (less than additive
of Pb associations to
effect observed)
other metals or
genes
Interactions between BLL and
Niqa et al. (2016)
genes that regulate
Roonev et al. (2018)
neurodevelopmental processes
observed.
Choi et al. (2020)
Association observed in
Ji et al. (2018)
Mean: 2.2 |jg/dL (<4 yr of age)
prospective study of clinical
ADHD diagnosed before age 6,
with adjustment for parental
education and SES.
No association found with
Wasserman et al. (2001)
Blood Pb: Mean 7.2 |jg/dL for lifetime (to age 4-5 yr)
ratings of attention problems in
avg
children ages 4-5 yr in whom
ratings may be measured less
reliably.
Supporting evidence
Associations of concurrent BLL
Section 4.3.3.1, U.S. EPA (2013)
Concurrent (ages 5-7.5 yr) blood Pb: Means 5.0-5.4
from cross-sectional
with attention decrements,
pg/dL
studies
impulsivity, and hyperactivity in
children ages 5-7.5 yr. Some
populations had high prenatal
drug or alcohol exposure.
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Rationale for
Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with Effects0
Epidemiologic
studies help rule out
chance, bias, and
confounding with
reasonable
confidence
Most prospective and some
cross-sectional studies found
associations with adjustment for
SES, maternal education, and
parental caregiving quality
(HOME score). Some also
considered parental IQ,
smoking, birth outcomes. A few
considered substance abuse,
nutritional factors, and family
history of psychiatric disorders.
Studies had population-based
recruitment with moderate to
high follow-up participation not
conditional on blood or tooth Pb
level.
Section 4.3.3.1, U.S. EPA (2013)
HOME score:
Liu etal. (2014b)
Fruh etal. (2019)
Family history of psychiatric disorders:
Choi etal. (2016)
SES and parental education:
Horton et al. (2018)
Ruebner et al. (2019)
Winter and Sampson (2017)
Boucher et al. (2012b)
Desrochers-Couture et al. (2019)
Consistent evidence
in animals with
relevant exposures
Several studies report increased
open-field activity in rodents
following developmental Pb
exposure, consistent with
hyperactivity.
Rodriques et al. (1996)
Moreira et al. (2001)
De Marco et al. (2005)
Duan etal. (2017)
Blood Pb: 19-28 |jg/dL in mice with lactational
exposure (tested PND 15-19); 10-29 |jg/dL in rats
with lactational exposure (tested PND 14-23)
Evidence from
lifetime Pb exposure
in nonhuman
primates suggests
that Pb produces
attention
decrements, which
supports the findings
in humans
Lifetime Pb exposure in
nonhuman primates was
reported to increase
distractibility in a spatial
discrimination task.
Gilbert and Rice (1987)
Blood Pb: 15-25 pg/dL
Evidence from lifetime Pb
exposure in nonhuman primates
suggests that Pb increased
perseveration and errors of
commission in a spatial
discrimination reversal task
Rice (1990)
Rice and Gilbert (1990b)
Gilbert and Rice (1987)
Blood Pb; 15-36 pg/dL
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Rationale for
Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with Effects0
Evidence describes
Both in vitro and in vivo Section 3.3
biologically plausible
evidence suggests that Pb
pathways
exposure may influence brain
development,
neurotransmission, connectivity,
neuronal integrity, all of which
may underlie the observed
alterations in externalizing
behaviors.
ADHD = attention deficit/hyperactivity disorder; avg = average; BLL = blood lead level; CKD = chronic kidney disease; Hg = mercury; HOME = Health Outcomes and Measures of the
Environment; mo = month(s); Pb = lead; PND = postnatal day; SES = socioeconomic status; yr = year(s).
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs CU.S. EPA. 20151.
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or inconsistencies.
References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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3.5.3 Externalizing Behaviors: Conduct Disorders, Aggression, and Criminal
Behavior in Children, Adolescents, and Young Adults
There are two domains of conduct disorders that are considered in the ISA: undersocialized
aggressive conduct disorder, and socialized aggressive conduct disorder (Whitcomb and Merrell, 2012).
As discussed in the 2013 Pb ISA (U.S. EPA, 2013), these domains are combined in this assessment
because they cannot be disentangled based on the available epidemiologic literature. This section also
considers evidence for criminal offenses, which are associated with conduct disorders (U.S. EPA, 2013).
Although not described explicitly as a part of either domain of conduct disorders, evidence for criminal
offenses is reviewed with conduct disorders because conduct disorders can be predictors of subsequent
delinquency and criminality (Soderstrom et al., 2004; Babinski et al., 1999; Pajer, 1998).
The evidence reviewed in the 2013 Pb ISA is sufficient to conclude that a "causal relationship is
likely to exist" between Pb exposure and conduct disorders in children and young adults (U.S. EPA,
2013). Prospective studies consistently indicated that earlier childhood (e.g., age 30 months 6 years) or
lifetime average (to age 11-13 years) BLLs or tooth Pb levels (shed between ages 6-8 years and typically
measured in dentin, which reflects prenatal and/or child Pb exposure depending on the tooth layer
analyzed, see Section 2.3.4.1.) were associated with criminal offenses in young adults aged 19-24 years,
and with higher parent and teacher ratings of behaviors related to conduct disorders in children ages 7-17
years (see Table 4-12 of (U.S. EPA, 2013) and (U.S. EPA, 2006)). Pb-associated increases in conduct
disorders were found in populations with mean BLLs of 7-14 (ig/dL. These associations were found
without indication of strong selection bias and with adjustment for SES, parental education and IQ,
parental caregiving quality, family functioning, smoking, and substance abuse. Supporting evidence was
provided by cross-sectional studies of children participating in NHANES, e.g., (Braun et al„ 2008), and a
meta-analysis of prospective and cross-sectional studies (Marcus et al„ 2010). In addition, there was
coherence across related measures of conduct problems in epidemiologic studies. Evidence for Pb-
induced aggression in animals was mixed, however, with increases in aggression found in some studies of
adult animals with gestational plus lifetime Pb exposure but not juvenile animals. The strongest evidence
for the 2013 causality conclusion was provided by prospective epidemiologic studies, with support from
cross-sectional studies of criminal offenses and ratings of behaviors related to conduct disorders.
Associations with lower BLLs that were not influenced by higher earlier Pb exposures as in older children
and adults were not well characterized, however.
Studies published since 2013 from both cohort and cross-sectional studies add to this evidence
base, which continues to support a "likely to be causal" relationship, as described in Table 3-4 and below.
The central tendency Pb levels, study-specific details, and selected effect estimates are highlighted in
Table 3-9E.
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3.5.3.1 Epidemiologic Studies of Conduct Disorders, Aggression, and Criminal
Behavior in Children and Adolescents
The 2013 Pb ISA describes cohort and cross-sectional studies demonstrating associations of Pb
exposure with behaviors related to conduct disorders, including criminal offenses (U.S. EPA, 2013).
Collectively, the evidence from prospective cohort studies indicated associations of aggressive, antisocial,
delinquent, and criminal behavior with biomarkers of Pb exposure. Cross-sectional studies also provided
evidence on these associations, though there is more uncertainty in data from this study design due to
limitations in assessing temporality.
Recent studies have evaluated associations between Pb and conduct disorders, aggressive
behavior, and other related measures of externalizing behavior. Most of these studies were prospective
cohort studies (Tlotlcng et al„ 2022; Desrochers-Couture et al., 2019; Reuben et al., 2019; Beckwith et
al„ 2018; Nkomo et al„ 2018; Nkomo et al., 2017; Liu et al., 2014b; Sioen et al„ 2013; Boucher et al.,
2012b; Tatsuta et al., 2012). Several utilized self-report tools (e.g., Youth Self-Report [YSR], Buss-Perry
Aggression Questionnaire [BPAQ], Psychopathic Personality Inventory [PPI], Antisocial Behavior
Interview) to assess aggression, violence, or other socio-behavioral problems among adolescents and
young adults aged 14-24 years. These studies reported central tendency BLLs at ages 6.5-13 years
ranging from 2.3 to 8 (ig/dL or mean bone Pb of 8.7 (ig/dL (Tlotlcng et al„ 2022; Desrochers-Couture et
al„ 2019; Beckwith et al„ 2018; Nkomo et al„ 2018; Nkomo et al., 2017). In analyses adjusted for most
key confounders, associations were observed for: physical violence (|3: 0.05; 95% CI: 0.04, 0.05) (Nkomo
et al., 2017); direct aggression (|3 [95% CI] comparing those with BLLs >10 (ig/dL to those with BLLs <5
Hg/dL: 0.43 [0.08, 0.78]) (Nkomo et al.. 2018); anger aggression (|3 = 0.25 [95% CI: 0.04, 0.37])
(Tlotleng et al.. 2022); and PPI (overall |3 = 0.22 [95% CI: 0.06, 0.38]; female |3 = 0.16 [95% CI: -0.05,
0.37]; male |3 = 0.22 [95% CI: -0.02, 0.47]) (Beckwith et al.. 2018) (Table 3-9E). Although the PPI
serves as a measure of psychopathic personality traits, psychopathy more generally includes behavioral
factors such as aggression and criminal conduct, in addition to personality traits. As discussed in the 2013
Pb ISA (U.S. EPA, 2013). an analysis of this same cohort reported associations between BLLs and
criminal and violent criminal arrests at ages 19-24 (Wright et al.. 2008). (Beckwith et al.. 2018) also
noted that BLLs were associated with volumetric reductions in gray matter in the frontal lobe and white
matter in several brain regions. Considered together, these studies provide support for an association
between childhood Pb exposure and psychopathy in adolescents and young adults that may stem from
changes in brain morphology. In addition to these studies examining total effects, one prospective study
of Pb and self-reported behavioral outcomes conducted mediation analyses and reported an association
between Pb and adolescent externalizing behavior mediated through child externalizing behavior (|3: 0.18,
95% CI: 0, 0.36) (Desrochers-Couture et al., 2019). There was also evidence of a small but imprecise
direct effect, though there was likely limited power to detect a direct effect given the small sample size
and correlation between child and adolescent externalizing behavior. The observed association between
BLLs and adolescent externalizing behavior is at least partially mediated through child externalizing
behavior in this study population.
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Other prospective cohort studies used observer assessments (e.g., parent or teacher ratings) to
assess conduct disorders, aggression, and related behaviors among children with central tendency blood
or cord blood Pb ranging from 0.4 to 14.3 (ig/dL (Fruh et al.. 2019; Ruebner et al.. 2019; Liu et al..
2014b; Sioen et al.. 2013; Boucher et al.. 2012b; Tatsuta et al.. 2012). Some of these studies focused on
the prenatal period as a potentially sensitive period of exposure: three of these studies evaluated Pb levels
in cord blood (Sioen et al.. 2013; Boucher et al.. 2012b; Tatsuta et al.. 2012) and one evaluated second
trimester maternal BLLs (Fruh et al.. 2019). All but two (Ruebner et al.. 2019; Boucher et al.. 2012b) of
these studies evaluated the outcome in children with mean age <8 years. These analyses reported
generally null associations (Table 3-9E). It is possible that behavioral ratings are less reliable at younger
ages or that these outcomes manifest at later ages (Blair. 2001).
There were also some cross-sectional evaluations of blood Pb and behavioral problems (e.g.,
aggression, oppositional, externalizing, antisocial) covering children with central tendency BLLs ranging
from 0.7 to 11.08 (ig/dL (Liu et al.. 2022b; Desrochers-Couture et al.. 2019; Reuben et al.. 2019; Barg et
al.. 2018; Rodrigucs et al.. 2018; Boucher et al.. 2012b; Naicker et al.. 2012; Nigg et al.. 2010). These
analyses utilized a mix of self-report and observer assessment tools to evaluate the outcomes of interest in
children aged 6-13 years. Positive associations were reported in most studies, including for child
externalizing behavior (|3: 0.23; 95% CI: 0.08, 0.38) and child oppositional defiant and conduct disorder
(OD/CD) (|3: 0.37; 95% CI: 0.06, 0.69) (Desrochers-Couture et al.. 2019); "attacking people" (boys only,
see table for unstandardized estimate; (Naicker et al.. 2012)); teacher-reported aggressive and rule-
breaking behavior (referred to in the paper as "externalizing behavior") (log-transformed concurrent Pb |3:
0.14; 95% CI: 0.01, 0.26) (Boucher et al.. 2012b); parent-reported externalizing composite (|3 for SD
increase in symptoms scores per SD increase in loglO transformed BLL = 0.21 [95% CI: 0.05, 0.37]) and
oppositional behavior (|3 for SD increase in symptoms scores per SD increase in log 10 transformed BLL
= 0.09 [95% CI: -0.09, 0.27]) (Nigg et al.. 2010); antisocial behavior (|3 = 0.02 [95% CI: 0.00, 0.04])
(Reuben et al.. 2019); and antisocial/aggressive behavior factor (Parent-reported |3 = 0.20 [95% CI: 0.05,
0.34]; Child-reported |3 = 0.20 [95% CI: 0.04, 0.35]) (Liu et al.. 2022b). Both of the null studies evaluated
the outcome in groups of children that included individuals aged <8 years (Barg et al.. 2018; Rodrigues et
al.. 2018); it is possible that behavioral ratings are less reliable at younger ages or that these outcomes
manifest at later ages (Blair. 2001). More research is needed to disentangle these issues. Overall, cross-
sectional studies were less of a consideration in drawing conclusions on the effects of Pb, given their
inherent limitations with regard to temporality (given exposure assessment using BLLs).
Among studies published since the 2013 Pb ISA, there were also evaluations of the association
between Pb and suspensions, arrests, juvenile delinquency, and crime (including violent crime) (Wright et
al.. 2021; Emer et al.. 2020; Becklev et al.. 2018; Boutwell et al.. 2017; Amato et al.. 2013). The strongest
evidence comes from three prospective studies. Using data from the CLS on multiple measures of BLLs
(from prenatal to age 6 years; mean = 14.4 (ig/dL) and arrests from ages 18-33, (Wright et al.. 2021)
observed numerous positive associations, including for adult arrests (RR =1.01 [95% CI: 1.00, 1.03]),
lifetime arrests (RR= 1.02 [95% CI: 1.00, 1.03]), arrests for violent crime (RR= 1.02 [95% CI: 0.99,
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1.04]), and arrests for drug crime (RR= 1.03 [95% CI: 1.01, 1.061) (Wright et al.. 2021). In a cohort
based in Milwaukee, Wisconsin, (Emer et al.. 2020) reported that elevated mean and peak BLL prior to
age 6 was associated with increased risk of firearm violence perpetration (RR for mean BLL = 1.03 [95%
CI: 1.02, 1.04]; RR for peak BLL = 1.02 [95% CI: 1.01, 1.02] (Emer et al.. 2020). Additionally, (Amato
et al.. 2013) reported that Pb exposure during the first 3 years of life (based on a BLL >10 (ig/dL and <20
(ig/dL) increased the odds of school suspensions in fourth grade, compared with those without Pb
exposure during the first three years of life (based on a BLL <5 (ig/dL) (OR: 2.66; 95% CI: 2.12, 3.32)
(Amato et al.. 2013).
Supporting evidence comes from a prospective study with limitations that affect interpretation
and confidence as well as one ecologic study. Criminal offending, comprising both criminal conviction
and self-report offending, was evaluated in a prospective cohort study based in Dunedin, New Zealand in
which the mean 11-year-old BLL was 11.01 (ig/dL (Becklev et al.. 2018). In sex-adjusted analyses of
convictions, the authors reported that increased childhood BLL was associated with increased odds of at
least one nonviolent criminal conviction forages 15-38 years (OR: 1.05; 95% CI: 1.00, 1.10). However,
sex-adjusted analyses of other criminal conviction endpoints (e.g., any criminal conviction, recidivistic
conviction, one-time conviction, violent offense) were inconclusive (see Table 3-9E). While this study
had extensive follow-up (27 years) and both subjective and objective measures of the outcome, the
limited adjustment for potential confounders is a concern. Analyses were adjusted for sex, age was
controlled in the study design, and SES was evaluated as a potential confounder but determined not to be
associated with BLL; however, important covariates (i.e., parental IQ or education and HOME score)
were not considered, leaving open the possibility of residual confounding. In an ecologic study of 106
census tracts in St. Louis, Missouri, United States, (Boutwell et al.. 2017) reported that a 1% increase in
the proportion of elevated blood Pb tests (>5 (ig/dL) among children within a census tract was associated
with increased RRs for firearm crimes (RR: 1.03; 95% CI: 1.03, 1.04), assault crimes (RR: 1.03; 95% CI:
1.02, 1.03), robbery crimes (RR: 1.03;95%CI: 1.02, 1.04), and homicides (RR: 1.03;95%CI: 1.01,
1.04). The association with rape was inconclusive (RR: 1.01; 95% CI: 0.99, 1.03). While ecologic studies
can be useful for hypothesis generation and understanding patterns among groups, the lack of control for
individual-level confounding factors in such studies leaves concern for risk of bias.
3.5.3.1.1 Summary
Overall, recently published epidemiologic studies support the findings from the previous ISA.
The strongest evidence published since 2013 comes from prospective cohort studies of 1) self-reported
conduct and aggression-related outcomes (Tlotlcng et al.. 2022; Desrochers-Couture et al.. 2019;
Beckwith et al.. 2018; Nkomo et al.. 2018; Nkomo et al.. 2017). and 2) external measures of delinquency
(e.g., criminal arrests, school suspensions) (Wright et al.. 2021; Amato et al.. 2013). These studies
evaluated outcomes among individuals aged 7-33 years in relation to earlier (or cumulative) Pb levels.
BLLs were <10 (ig/dL in the studies of self-reported conduct and aggression-related outcomes and higher
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in studies of external measures of delinquency (e.g., (Wright et al.. 2021); mean 14.4 (ig/dL). These
studies controlled for most relevant confounders, and the prospective study design inherently ensured
appropriate temporality between the exposure and outcome. Additional supporting evidence comes from
cross-sectional studies using either self-report or observer-reported outcome measures among individuals
aged 6-13 years with concurrent BLLs ranging from 0.7-11.08 (ig/dL (Liu et al.. 2022b; Desrochers-
Couture et al.. 2019; Reuben et al.. 2019; Boucher et al.. 2012b; Naicker et al.. 2012; Nigg et al.. 2010).
3.5.3.2 Toxicological Studies of Aggression
There are no recent PECOS-relevant studies examining the relationship between Pb exposure and
aggression. Available toxicological studies of aggression were described in the 2006 Pb AQCD (U.S.
EPA, 2006) and 2013 Pb ISA (U.S. EPA, 2013). The evidence supported effects of Pb exposure on
changes in social behavior of rodents and nonhuman primates. In animals, the social behavior most
comparable to conduct disorders in children is aggression; however, the effects of Pb on aggression in
animals were inconsistent. In animals, aggression was assessed as threats, attacks, bites, chases, and
offensive posture in encounters with other animals. Pb exposure was found to have no effect on
aggression in some studies as well as to decrease and increase aggression in others. Pb exposure generally
was not found to affect aggression in juvenile animals; however, increased aggression was found in adult
animals with high concentrations of gestational plus postnatal dietary Pb exposure. Recent PECOS-
relevant studies have not further examined the effects of Pb on aggression. Additional reported effects on
social behaviors described in the 2006 Pb AQCD (U.S. EPA, 2006) and 2013 Pb ISA (U.S. EPA, 2013)
included Pb-induced increases in social and sexual investigation, as indicated by sniffing, grooming,
following, mounting, and lordosis behavior. Despite the limited new evidence, observations for Pb-
induced changes in aggression in animals provide support for associations of altered aggression outcomes
in children. Furthermore, many of the more general overt nervous system toxicology studies discussed in
Sections 3.4.2 and 3.3 assessed a variety of endpoints, including brain structural changes and
neurotransmitter analysis, that can contribute to understandings of the mechanistic underpinning of
observed behavioral changes providing additional biological plausibility.
3.5.3.3 Relevant Issues for Interpreting the Evidence Base
3.5.3.3.1 Concentration-Response Function
The evidence base for this outcome is more limited compared with that for cognitive deficits, and
the shape of the C-R function cannot be determined from available studies. However, it is important to
highlight that in studies reviewed for the 2013 Pb ISA, effects were observed at central tendency BLLs of
5-10 (ig/dL (Nigg et al.. 2008; Wright et al.. 2008; Chiodo et al.. 2007; Wasserman et al.. 2001) and <5
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(ig/dL (Braun et al.. 2008). Among studies published since 2013, effects on conduct disorder, aggression,
and crime were observed at central tendency BLLs of 5-10 (ig/dL (Tlotlcng et al.. 2022; Beckwith et al..
2018; Nkomo et al.. 2018; Nkomo et al.. 2017; Naicker et al.. 2012) as well as at central tendency BLLs
<5 (ig/dL (Liu et al.. 2022b; Desrochers-Couture et al.. 2019; Boucher et al.. 2012b; Nigg et al.. 2010).
However, it should be noted that there is less confidence in these studies of BLLs <5 (ig/dL as they were
all cross-sectional analyses (Liu et al.. 2022b; Desrochers-Couture et al.. 2019; Boucher et al.. 2012b;
Nigg et al.. 2010). Further work is needed to better understand whether the potential effects of Pb on this
outcome persist at BLLs <10 (ig/dL.
3.5.3.3.2 Potentially At-Risk Populations
Sex
The 2013 Pb ISA identified one study that evaluated the role of sex as an at-risk factor. Wright et
al. (2008) examined early life BLLs and criminal arrests in adulthood and reported that risks attributable
to Pb exposure were greater among males than females (Wright et al.. 2008).
Several new studies evaluated the role of sex as an at-risk factor through sex-stratified analyses of
Pb exposure and conduct disorders. The results were generally inconclusive regarding sex as an at-risk
factor. In a prospective study of blood Pb concentrations at 3-5 years and teacher-rated behavioral
problems at age 6 years, sex-stratified results were similar to non-stratified results, with null associations
for conduct disorder and aggression-related outcomes (Liu et al.. 2014b). In a prospective study of Pb
exposure and self-reported aggressive behavioral characteristics, associations for some outcomes (e.g.,
"attacks people") were observed in boys (but not observed or reported for girls); the authors suggested
this may be due to lower BLLs in girls compared with boys (Naicker et al.. 2012).
In a cross-sectional study of first grade children (mean 6.7 years) and teacher-rated behavioral
problems, sex-stratified results were generally null and similar to the non-stratified results. However,
some analyses indicated stronger associations among females (e.g., Behavioral Regulation Index (PR
[95% CI]: girls = 1.03 [1.00, 1.05]; boys = 0.99 [0.97, 1.01]), though the sample size was limited (n = 83
for girls) (Barg et al.. 2018). The authors suggested these results could be explained by teacher
expectations and perceptions of girls compared with boys, with effects on girls being more noticeable due
to gender norms and expectations rather than greater susceptibility to Pb exposure (Barg et al.. 2018).
Finally, in a study of BLLs measured at age 6.5 years and PPI between ages 19 and 24, sex-
stratified models indicated stronger associations in males, though associations were also present in
females (Beckwith et al.. 2018). Sex-stratified analyses indicated that Pb-associated gray matter volume
loss was only present in females, while Pb-related white matter loss was more widespread in males,
including an overlap in frontal white matter loss associated with both PPI scores and BLLs (Beckwith et
al.. 2018).
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Pre-existing Conditions
One study evaluated the association between Pb (median BLL 1.2 (ig/dL) and aggression/conduct
problems among children with CKD, a population at elevated risk of neurocognitive dysfunction (Gerson
et al., 2006; Gipson et al., 2004). No associations were observed (Ruebner et al.. 2019).
3.5.3.3.3 Confounding
The 2013 Pb ISA described multiple factors that influence conduct disorder and related outcomes
including sex, race, SES, parental education, parental IQ, and quality of the caregiving environment (i.e.,
HOME score) (U.S. EPA, 2013). These risk factors are often correlated with blood, tooth, and bone Pb
levels, and thus, are considered as potential confounding factors in epidemiologic analyses. As noted in
the 2013 Pb ISA, no single method to control for potential confounding is without limitation, and there is
potential for residual confounding by unmeasured factors. However, consistency of results across studies
utilizing different approaches to control for confounding can increase confidence across the body of
evidence.
Recent studies demonstrate associations between Pb exposure and conduct disorder after
controlling for different combinations of the aforementioned key covariates as well as additional relevant
covariates. However, it should be noted that in the current evidence base, the vast majority of studies that
identified associations did not specifically adjust for HOME score (Liu et al„ 2022b; Tlotlcng et al., 2022;
Desrochers-Couture et al., 2019; Reuben et al., 2019; Barg et al., 2018; Beckley et al„ 2018; Nkomo et
al., 2018; Rodrigues et al., 2018; AbuShady et al., 2017; Boutwell et al., 2017; Nkomo et al., 2017; Liu et
al„ 2014b; Amato et al„ 2013; Sioen et al., 2013; Boucher et al„ 2012b; Naicker et al., 2012; Tatsuta et
al., 2012; Nigg et al., 2010). Yet, most of these studies did adjust for other potentially related covariates
such as social adversity, house crowding, family violence, and SES, which mitigates some of the concern
about residual confounding due to exclusion of HOME score. Additionally, as highlighted in the previous
ISA, a meta-analysis by Marcus et al. indicated that the lack of adjustment for variables such as SES or
HOME score does not warrant limiting inferences from a particular study (U.S. EPA, 2013; Marcus et al.,
2010).
When there is uncertainty in epidemiologic evidence due to potential confounding, it is often
helpful to consider associated toxicological data. Aggressive behavior in rodents is mediated by several
brain regions, including the hypothalamus, prefrontal cortex, dorsal raphe nucleus, nucleus accumbens,
and olfactory system (Takahashi and Miczek, 2014) along with other neurochemical systems including
neurotransmitters, neuropeptides, and neuromodulators (i.e., serotonin, dopamine, vasopressin, oxytocin,
testosterone, estrogen, corticotrophin releasing factor, opioids, neuronal nitric oxidate synthase, and
monoamine oxidase A) (Takahashi and Miczek, 2014). Pb-induced changes on many of these
neurochemical endpoints has been reported and are described in Section 3.3, which lends some limited
yet relevant biological plausibility from the animal evidence without influence of potential confounding
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factors. While no new studies on Pb-induced aggressive behavior in mammals were identified with BLLs
of relevance to this ISA, the previous experimental animal studies support the evidence described in the
2013 Pb ISA.
3.5.3.3.4 Lifestages
Environmental exposures during critical lifestages spanning from childhood into adolescence can
affect key physiological systems that orchestrate brain development and plasticity (see Section 3.5.1.6.4).
Epidemiologic evidence assessed in the 2013 Pb ISA indicated associations of earlier childhood blood or
tooth Pb levels with behaviors related to conduct disorders in adolescents or adults (Fergusson et al..
2008; Wright et al.. 2008); however, these epidemiologic studies did not examine adult BLLs, thus the
relative influence of adult Pb exposure cannot be ascertained.
Recent studies observed associations of Pb exposure assessed via blood, cord blood, or bone
between delivery and age 13 years with outcomes evaluated among children, adolescents, and young
adults aged 7-33 years (Liu et al.. 2022b; Tlotleng et al.. 2022; Wright et al.. 2021; Desrochers-Couture et
al.. 2019; Reuben et al.. 2019; Beckwith et al.. 2018; Nkomo et al.. 2018; Nkomo et al.. 2017; Amato et
al.. 2013; Naicker et al.. 2012). Evidence published since 2013 is weaker for exposures that occur during
the prenatal period (Fruh et al.. 2019; Sioen et al.. 2013; Tatsuta et al.. 2012) and for most studies
assessing outcomes prior to the age of 8 years (Fruh et al.. 2019; Liu et al.. 2014b; Sioen et al.. 2013;
Tatsuta et al.. 2012). It is possible that outcome assessment tools that measure conduct disorder and
related aggressive traits are less reliable in this age group, aggressive patterns have not yet stabilized, or
the particular type of aggression associated with Pb exposure does not manifest until later years (Blair.
2001). Overall, Pb exposure during lifestages spanning childhood and into adolescence may confer risk
for conduct disorders and related outcomes.
3.5.3.3.5 Public Health Significance
The global prevalence of conduct disorders in 2019 was estimated to be 40.1 million (95% CI: 29
million, 52 million), with the highest burden experienced by individuals 0-14 years of age (GBP 2019
Mental Disorders Collaborators. 2022). Early life conduct disorders and other "antisocial behaviors" are
an important public health issue due to their persistence within an individual (Lvnam et al.. 2009). their
costs (both social and economic) to society (Sumner et al.. 2015; Mccollister et al.. 2010). and their
association with risk-taking behaviors, comorbid mental health conditions, and premature mortality
(Reves. 2015; Maughan et al.. 2014; Glenn et al.. 2013). For example, in one recent study based in New
Zealand, children with conduct problems accounted for 9.0% of the population but 53.3% of convictions,
15.7% of emergency department visits, 20.5% of prescription fills, 13.1% of injury claims, and 24.7 % of
welfare benefit months (Rivenbark et al.. 2018).
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3.5.3.4 Summary and Causality Determination: Conduct Disorders, Aggression, and
Criminal Behavior
The 2013 Pb ISA concluded that the relationship between Pb exposure and conduct disorders was
"likely to be causal" (U.S. EPA. 2013). This causality determination was primarily based on
epidemiologic evidence. In particular, prospective cohort studies provided key evidence of the association
between blood or tooth Pb levels and 1) parent or teacher ratings of delinquent, aggressive, and antisocial
behavior (Chandramouli et al„ 2009; Dietrich et al„ 2001; Burns et al„ 1999), and 2) criminal offenses
(Fergusson et al., 2008; Wright et al., 2008) in children and adolescents across diverse locations.
Supporting evidence was provided by cross-sectional studies of these outcomes (Braun et al„ 2008;
Chiodo et al., 2007).
Recent epidemiologic studies support the findings from the previous ISA. The strongest evidence
published since the 2013 Pb ISA comes from prospective cohort studies of 1) self-reported conduct and
aggression-related outcomes (Tlotlcng et al„ 2022; Desrochers-Couture et al„ 2019; Beckwith et al„
2018; Nkomo et al., 2018; Nkomo et al„ 2017), and 2) external measures of delinquency (e.g., criminal
arrests, school suspensions) (Wright et al„ 2021; Amato et al„ 2013). BLLs were <10 (ig/dL in studies of
self-reported conduct and aggression-related outcomes and higher in studies of external measures of
delinquency (e.g., (Wright et al„ 2021); mean 14.4 (ig/dL). These studies controlled for most relevant
confounders, and the study design inherently ensured appropriate temporality between the exposure and
outcome. Additional supporting evidence comes from cross-sectional studies using either self-report or
observer-reported outcome measures among individuals aged 6-13 years with concurrent BLLs ranging
from 0.7-11.08 (ig/dL (Liu et al., 2022b; Desrochers-Couture et al., 2019; Reuben et al„ 2019; Boucher et
al„ 2012b; Naicker et al„ 2012; Nigg et al., 2010). Although the evidence generally suggests positive
associations, null results may be explained by age at outcome or exposure. For example, many studies
with null associations evaluated the outcome in groups of children that included individuals <8 years of
age. It is possible that behavioral ratings are less reliable among this younger age group and/or abnormal
behaviors do not manifest until later in childhood. It should also be noted that both studies focusing
exclusively on newborn exposure (i.e., measurement of Pb in cord blood) were null, which potentially
indicates that the prenatal period may not be a relevant sensitive period of exposure for this outcome.
Studies that provide information on sensitive periods of exposure are limited.
Despite the growing epidemiologic evidence, the central uncertainty present in the 2013 Pb ISA
database remains: there is limited and inconsistent evidence from animal toxicological studies. Available
toxicological studies of aggression were described in the 2006 Pb AQCD (U.S. EPA, 2006) and 2013 Pb
ISA (U.S. EPA, 2013). No new PECOS-relevant studies examining the relationship between Pb exposure
and aggression have been reported. Despite the lack of new PECOS-relevant studies, Pb-induced changes
on many neurochemical endpoints that contribute to aggressive behaviors have been reported and are
described in Section 3.3, which lends biological plausibility from the animal evidence.
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In summary, there is sufficient evidence to conclude that there is likely to be a causal
relationship between Pb exposure and conduct disorders, aggression, and criminal behavior. This
causality determination is based on positive associations observed across various populations and based
on multiple outcome assessment approaches at relevant Pb exposure levels across recently published
prospective and cross-sectional epidemiologic studies. However, limitations remain in the animal
toxicology database, given the inconsistent evidence described in the 2013 Pb ISA and the lack of
relevant studies published since then. Yet, biological plausibility for these associations is supported by
human evidence linking early life Pb exposure to later life volumetric reductions in gray matter in the
frontal lobe and white matter in several brain regions (Bcckw ith et al.. 2018) and experimental animal
studies demonstrating Pb-induced changes on neurochemical endpoints relevant to this set of outcomes.
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Table 3-4 Summary of evidence for a likely to be causal association between Pb exposure and conduct
disorders, aggression, and criminal behavior in children and adolescents
Rationale for Causality Key Evidence* References* Pb Biomarker Levels Associated
Determination3 y with Effects0
Consistent results from
epidemiologic studies
with relevant blood or
bone Pb levels,
adequate control of
relevant confounders
Evidence from prospective studies
(demonstration of a temporal sequence)
using self-report measures of aggressive or
related externalizing behavior among
individuals ages 14-24 yr in relation to
earlier average blood Pb or bone Pb
Nkomo etal. (2017)
Nkomo etal. (2018)
Beckwith etal. (2018)
Tlotlenq et al. (2022)
Blood Pb: age 6.5-13 mean = 5.6-
pg/dL
Bone Pb: age 9 mean = 8.7 |jg/g
Evidence from prospective studies Wright et al. (2021) Blood Pb: prenatal to age 6 mean =
(demonstration of a temporal sequence) of Amato et al (2013) >10 [jg/dL
arrests (ages 18-33 yr) and suspensions
(ages 9-10 yr) in relation to earlier average
blood Pb
Supporting evidence from cross-sectional
studies using both self-report and observer-
reported measures of aggressive or
externalizing behavior among individuals
ages 6-13 yr
Desrochers-Couture et al. (2019)
Naickeret al. (2012)
Liu et al. (2022b)
Boucher et al. (2012b)
Nigg et al. (2010)
(Reuben et al.. 2019)
Blood Pb (concurrent): age 6-13
mean = 0.7-11.08 pg/dL
Blood Pb: age 11 geometric mean =
2.3 |jg/dL
indicating indirect association of BLL on
adolescent externalizing behavior via child
externalizing behavior
Supporting evidence from a mediation Desrochers-Couture et al. (2019)
analysis from a prospective cohort study
Supporting evidence from a prospective Beckwith et al. (2018) Blood Pb: age 6.5 mean = 8.0 |jg/dL
study demonstrating association between
early life BLL and volumetric reductions in
gray matter in the frontal lobe and white
matter in several brain regions (mean age
26.8 yr)
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Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated
with Effects0
Most studies had sufficient adjustment for
relevant confounders. While most did not
adjust for HOME score specifically, they did
consider other related variables, such as
income, parental IQ, parental education,
SES, and/or neighborhood safety
Evidence strongest for outcomes assessed
among children >9 yr
Experimental animal Supporting evidence from animals exposed U.S. EPA (2013)
studies with relevant prenatally and postnatally
exposures provide
coherence and help rule
out chance, bias, and
confounding with
reasonable confidence
Biological plausibility Changes in key brain regions and U.S. EPA (2013)
demonstrated neurochemical systems implicated in
behavioral changes.
BLL = blood lead level; HOME = Health Outcomes and Measures of the Environment; IQ = intelligence quotient; SES = socioeconomic status; yr = year(s).
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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3.5.4
Internalizing Behaviors: Anxiety and Depression in Children
The evidence evaluated in the 2013 Pb ISA was sufficient to conclude that a "causal relationship
was likely to exist" between Pb exposure and internalizing behaviors in children (U.S. EPA, 2013).
Prospective studies in a few populations found associations of higher lifetime average blood Pb (mean:
-14 (ig/dL) or childhood tooth Pb (shed between ages 6-8 years and generally reflecting prenatal or early
child Pb exposure depending on the tooth layer analyzed, see Section 2.3.4.1) levels with higher parent
and teacher ratings of internalizing behaviors such as withdrawn behavior and symptoms of depression
and anxiety in children aged 8-13 years. There was no strong indication of biased reporting of behaviors
for children with higher BLLs. The few cross-sectional associations in populations with mean concurrent
BLLs of ~5 (ig/dL were inconsistent. Pb-associated increases in internalizing behaviors were found with
adjustment for maternal education and SES-related variables. Consideration for potential confounding by
parental caregiving quality was inconsistent. Despite some uncertainty in the epidemiologic evidence, the
biological plausibility for the effects of Pb on internalizing behaviors was provided by a small number of
experimental animal study findings with dietary lactational Pb exposure, with some evidence at BLLs
relevant to humans. Additional toxicological evidence demonstrating Pb-induced changes in the HPA axis
and dopaminergic and gamma-aminobutyric acid (GABA) systems provided additional support. Overall,
the strongest evidence was from prospective studies in a few populations of children and the coherence
with evidence from a small number of experimental animal studies with relevant Pb exposures. Some
uncertainty related to potential confounding by parental caregiving quality remained.
Measures of central tendency for Pb biomarker levels used in each study, along with other study-
specific details, including study population characteristics and select effect estimates, are highlighted in
Table 3-10E (Epidemiologic Studies) and Table 3-7T (Toxicological Studies). An overview of the recent
evidence is provided below. Overall, recent studies generally support findings from the 2013 Pb ISA.
3.5.4.1 Epidemiologic Studies of Internalizing Behaviors in Children
Several epidemiologic studies evaluated in the 2013 Pb ISA linked biomarkers of Pb exposure in
children with internalizing behaviors characterized by directing feelings and emotions inward, i.e.,
withdrawn behavior, symptoms of depression, fearfulness, and anxiety. These studies did not clearly
indicate that Pb exposure affected a particular domain of internalizing behaviors, i.e., withdrawn
behavior, somatic symptoms, anxiety, and depression. However, a consistent pattern of associations with
BLLs was observed across ages and across multiple internalizing behaviors. The strongest evidence was
provided by prospective studies conducted across multiple locations, i.e., Boston, Port Pirie, Australia,
and Yugoslavia (Wasserman et al.. 2001; Burns et al.. 1999; Wasserman et al.. 1998; Bellinger et al..
1994b). Collectively, these studies found associations between internalizing behaviors in children (ages
3-13 years) and Pb levels based on cord blood, concurrent blood (age 3 years), lifetime average blood,
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and teeth. Moderate to high follow-up rates in most studies increased confidence that selection bias did
not explain the pattern of associations observed in the studies. Factors, which were well documented to be
correlated with both Pb exposure and internalizing behaviors, including SES, parental caregiving quality
(speculated to mediate the potential correlation between parental psychopathology and Pb exposure), and
parental education, were considered as potential confounders across most studies. Although internalizing
behaviors are likely to have a strong familial component, the available evidence did not support parental
psychopathology as a direct confounder of the child Pb-internalizing behavior association. Studies that
included both teacher and parent ratings were emphasized. The most common instrument used to assess
internalizing behaviors was CBCL. Summary scores for internalizing behaviors, associated syndromes,
and DSM-IV scales (e.g., anxiety and depression) can be derived using CBCL.
Recent studies also analyze the association of Pb exposure with internalizing behaviors assessed
using CBCL. Using community survey data from the Project on Human Development in Chicago
Neighborhoods (PHDCN), Winter and Sampson (2017) examined the relationship between average BLL
in childhood (6 years old or younger) and anxiety or depression in adolescence (mean age 17 years old).
These authors found a 0.09 SD (0.03, 0.16) increase in anxiety or depression score, after adjustment for
covariates including caregiver education and SES. Participants were originally enrolled in the mid-1990s
and a random sample of those continuing to participate in 1999 and 2002 was randomly selected for this
study, with 67% of those selected agreeing to participate. Liu et al. (2014b) examined the association of
early childhood blood Pb concentration (3, 4, or 5 years old) with both parent and teacher ratings of
internalizing behavior at age 6 using CBCL and C-TRF, respectively. The outcomes were modeled as
both continuous and dichotomous variables (i.e., clinically significant behavior problems with T-score
>60) and adjusted for potential confounders including parent's educational level, father's occupation, and
child IQ. The emotional reactivity syndrome component of the teacher-rated internalizing problem scale
and the DSM-IV oriented anxiety were associated with child BLL when scores were modeled as
continuous terms (|3 = 0.32 [95% CI: 0.06, 0.59] and |3 = 0.25 [95% CI: 0.02, 0.50], respectively). The
ORs were 1.10 (95% CI: 1.03, 1.18) for the association of child BLL with clinically significant teacher-
reported internalizing behavior and 1.10 (95% CI: 1.01, 1.19) for clinically significant anxiety problems.
The participation rate was 81% in this study. The mean BLL of the children in this study was 6.4 (ig/dL
and the study had a high participation rate and included both teacher and parent ratings of internalizing
behavior.
Joo et al. (2018) analyzed data from the MOCEH study, a Korean prospective birth cohort of
mother-child pairs that were followed for 5 years. Maternal (early and late pregnancy), cord, and multiple
postnatal blood Pb concentrations were measured, and internalizing behaviors were assessed by the parent
using the Korean-CBCL at age five. The interaction between Pb exposure and child sex was evaluated
with further model adjustment for covariates including maternal educational level, and SES. Late
pregnancy and cord BLLs were associated with increasing internalizing behavior ratings in boys (|3 = 2.55
[95% CI: 0.22, 4.88] and |3 = 2.44 [95% CI: -0.74, 5.63], respectively), while postnatal (ages 2 and 5)
BLL was associated with increasing internalizing behavior ratings in girls (|3 = 2.94 [95% CI: 0.36, 5.52]
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and |3 = 5.65 [95% CI: 0.5, 10.8]). A total of 579 women of the 1751 originally enrolled in the cohort
provided data for this study.
Recent studies also examined the association of Pb exposure with internalizing behaviors using
SDQ. SDQ includes five scales (i.e., peer relationship problems, hyperactivity, emotional problems,
conduct problems and prosocial behavior) with the results for emotional problems discussed in this
section. Fruh et al. (2019) studied mother-child pairs participating in Project Viva, a longitudinal birth
cohort in eastern Massachusetts. Maternal blood Pb concentration in erythrocytes was measured during
the second trimester of pregnancy and parents rated their child's behavior using the SDQ (see also
Sections 3.5.1, 3.5.2, 3.5.3) in mid-childhood (median 7.7 years). The associations (i.e., P coefficients)
with the parent- and teacher-rated emotional components of the SDQ were 0.30 (95% CI: 0.05, 0.55) and
0.07 (95% CI; -0.22, 0.35), respectively. A stronger association with the emotional component of the
SDQ for girls compared with boys was reported by parents (P = 0.52 [0.18, 0.86] for girls versus P = 0.17
[95% CI: -0.17, 0.50] for boys). Note that the higher scores on the emotional problem scale indicate
worse performance. Behavior assessments and maternal blood Pb measurements were available for fewer
than half of study participants; however, important confounders including HOME score, maternal IQ, and
parental education were considered in this study. Sioen et al. (2013) analyzed data from a birth cohort
(FLEHS I, 2002-2006) comprising mother-infant pairs born in the Netherlands. This study examined the
association of cord blood for 281 infants whose parents returned the SDQ (26.4% response rate). No
association of cord blood Pb concentration with emotional symptom score >5 was observed (OR: 0.90
[95% CI: 0.52, 1.55] per doubling of BLL on log-scale).
Rokoff et al. (2022) used Couriers' Parent and Teacher Ratings Scales (CPRS and CTRS) at age 8
years and the BASC-2 self-report of personality (SRP) at age 15 years to assess internalizing behaviors
among children enrolled in a birth cohort study in New Bedford, MA. This study examined the
association of cord blood Pb with internalizing behaviors and also considered exposure to
organochlorines (hexachlorobenzene, p.p'-dichlorodiphenyl dichloroethylene, polychlorinated biphenyls)
and Mn, which were also measured in cord blood. BKMR analysis indicated linear associations and no
interactions between cord Pb, Mn, and organochlorines. Cord blood Pb was positively associated with
BASC anxiety score at age 15 (P = 1.78 [95% CI: 0.58, 2.99] BASC-2 SRP anxiety score increase per
doubling Pb) but not with Couriers' anxious-shy score at age 8 years. Additionally, a positive association
of cord blood Pb with depression score at age 15 was observed (P = 0.79 [95% CI: -0.39, 1.97]). The
Connor's psychosomatic score was positively associated with cord Pb, and this association was stronger
in boys (P = 2.08 [95% CI: 0.07, 4.10]) than in girls (P = 0.48 [95% CI: -1.00, 1.97]). A total of 528 of
the original 788 (67%) mother-infant pairs participated in the 15-year follow-up. The models were
adjusted for SES, maternal age, smoking, seafood, alcohol intake during pregnancy, maternal IQ, quality
of parental caregiving, and child characteristics (sex, race/ethnicity, age at assessment).
Several additional studies used the BASC-2 to assess associations with Pb exposure. Rasnick et
al. (2021) designed a study to identify sensitive time windows of exposure to Pb in air. These authors
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controlled for concurrent BLL (age 12 years) in their analysis of the Cincinnati Study of Allergy and Air
Pollution study data. Air Pb exposure was estimated using validated land use regression models, and
behavioral outcomes, including depression and anxiety, were assessed using the BASC-2 administered at
age 12. Models were adjusted for community deprivation, residential greenspace, and ECAT, in addition
to concurrent BLL. Distributed lag models that predicted outcome responses based on current and past
(i.e., lagged) predicted air Pb exposures identified a sensitive window in late childhood for anxiety but not
depression (Figure 3-12). The sensitive time window is indicated by months when the estimated 95% CI
did not include the null value.
BASC-2 = Behavior Assessment System for Children; edf =effective degrees of freedom.
The solid lines show the predicted change in score and the gray shading indicates the 95% CIs.
Source: Rasnick et al. ('20211.
Figure 3-12 Associations of monthly airborne Pb exposure levels from birth to
age 12 with scores for anxiety and depression behaviors on the
Behavior Assessment System for Children.
Ruebner et al. (2019) evaluated the association between BLLs and attention among children with
CKD. Internalizing behavior symptoms were assessed using the parental rating scales of the BASC-2,
which includes a composite score for internalizing problems. Associations between BLL and behavioral
symptoms on BASC-2 did not persist in models that were controlled for potential confounders including
race, poverty, maternal education, and clinical factors related to CKD. The median BLL in this study was
1.2 (.ig/dL.
Two additional prospective studies examined the association Pb concentration in teeth and
toenails with internalizing behavior on the BASC; these studies relied on a low proportion of the original
cohort, however. Horton et al. (2018) analyzed data from the ELEMENT Project birth cohort in Mexico
City to determine the association of weekly tooth Pb concentration (prenatal through 1 year postnatal)
with BASC-2 scores assessed between 8 and 11 years old. Approximately 12% of the original cohort was
enrolled in this study. Participants differed with respect to child birth weight and maternal IQ. A 0.4-unit
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increase in anxiety score was associated with a log-transformed unit increase in tooth Pb concentration at
12 months, while no consistent pattern of association was observed with increased internalizing behavior
symptoms overall. Dohertv et al. (2020) followed children enrolled in the New Hampshire Birth Cohort
Study (NHBCS) to examine the association of toenail Pb concentration with parent-rated internalizing
behaviors on the BASC-2. Data were available for approximately 300 of the 2000 women enrolled in the
study. No consistent pattern of association between pre- or postnatal toenail Pb concentration was
observed with internalizing behaviors after adjustment for confounders, including parent education and
parent perception of the parent-child relationship.
3.5.4.1.1 Summary
The 2013 Pb ISA included several prospective studies with moderate to high participation rates
that controlled for potential confounders including SES, parental education, and quality of parental
caregiving. These studies found associations of higher lifetime average blood (mean: -14 (ig/dL) or
childhood tooth Pb levels with higher parent and teacher ratings of internalizing behavior on the CBCL in
children aged 8-13 years. Several recent longitudinal epidemiologic studies with high to moderate
participation rates, which relied on an expanded array of instruments to assess internalizing behaviors
(i.e., CBCL, SDQ, CPRS, CTRS, and BASC-2), reported associations with blood Pb concentration
(childhood average, prenatal, and postnatal BLLs <7 (.ig/dL). Several studies in children evaluated sex
(Rokoff et al.. 2022; Fruh et al.. 2019; Joo et al.. 2018) as an effect modifier. The majority of analyses
controlled for important potential confounders including the quality of parental caregiving (Rokoff et al..
2022; Fruh et al.. 2019) maternal education and SES (Rokoff et al.. 2022; Fruh et al.. 2019; Winter and
Sampson. 2017; Liu et al.. 2014b). No association with internalizing behaviors was observed for the
blood Pb of children with CKD or in prospective studies of Pb concentration in blood (Sioen et al.. 2013).
teeth (Horton et al.. 2018). or toenails (Dohertv et al.. 2020). which reported relatively low participation
rates. The limited number of studies that aimed to distinguish types of internalizing behaviors indicated
associations with the anxiety component (Rokoff et al.. 2022; Rasnick et al.. 2021).
3.5.4.2 Toxicological Studies of Anxiety and Depression
Evidence in the 2013 Pb ISA consistently supported increases in emotionality in Pb-treated
animals. Postnatal exposure to Pb in female Long-Evans rats, resulting in mean BLLs between 13 and 31
(ig/dL, increased disruption and frustration in response to errors and reward omission in discrimination
task trials (Beaudin et al.. 2007; Stanglc et al.. 2007). Pb-exposed female Rhesus macaques displayed
increased negative responses to repeated tactile stimuli (i.e., tactile defensiveness) during adolescence
(mean BLLs of 31 (ig/dL) (Moore et al.. 2008). Furthermore, decreased exploratory behaviors in the
open-field test were also reported in male Wistar rats following Pb exposure from gestation through
weaning (Souza Lisboa et al.. 2005). Additional evidence for increased anxiety-like behavior, evaluated
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via the elevated plus maze, was found in one study of postnatally exposed rats (mean BLLs 35 (ig/dL at
weaning) (Fox et al.. 2010); however, no significant effects were found in another study of postnatal Pb
exposure (Molina et al.. 2011). Inconsistent evidence for depression-like behaviors measured in the forced
swim test was also reported (Souza Lisboa et al.. 2005; Stewart et al.. 1996).
Tests of anxiety-like behavior (i.e., emotionality) in rodents are often designed to exploit the
approach-avoidance conflict. Rodents must balance their motivation to explore novel environments (to
gather food and resources) with the need to evade predators and other threats. The open-field test (OFT)
allows for observation of rodent behavior within a bare, brightly lit, open area. Decreases in measures of
exploration (e.g., rearing, sniffing) indicate a shift towards an anxiety-like phenotype, although some
metrics may also be affected by other factors such as decreased motor function, to varying degrees. Basha
et al. (2014) found that postnatal Pb exposure in male rats decreased rearing and sniffing in the OFT
between PND 45 and 18 months, well after exposure was terminated. Grooming was also decreased at
PND 45, 4 months, and 12 months, which may indicate an altered response to stress in comparison to
controls. The same study also utilized the hole board test as another method to evaluate rodents" interest
in exploration of a novel environment. Animals displayed anxiety-like behavior (i.e., decreases in head
dip count and head dip duration) between PND 45 and 18 months. A follow-up study evaluated male
Wistar rats using a prenatal Pb exposure paradigm that resulted in BLLs of 11 (ig/dL at PND 21 and
found decreased exploratory behaviors in both the OFT and hole board test between PND 21 and 4
months (Basha and Reddv. 2015). Decreases in head dipping behavior were also reported by Flores-
Montovaand Sobin (2015). who evaluated male and female C57BL/6 mice following exposure to Pb
from PND 0 to PND 28 that resulted in low BLLs (mean between 3 and 12 (ig/dL). Further analysis of
individual BLLs and head dipping behavior suggested a negative association (i.e., head dipping behaviors
decreased as BLLs increased).
Enhanced thigmotaxis (i.e., tendency to remain close to the walls of the arena) within the OFT is
also associated with an anxiety-like phenotype. Betharia and Maher (2012) reported that low dose Pb
treatment had no significant effects on the latency of rodents to enter the center of the arena at PND 24 or
PND 59. The Sprague Dawley rats used in this study were exposed to Pb through their mothers from
gestation until PND 20 and had a mean BLL of 9 (ig/dL at PND 2, which decreased to <1 (ig/dL when
behavior was assessed. Another recent study found that adolescent Pb exposure (between PND 24 and
PND 56) in male Sprague Dawley rats, resulting in mean BLLs of 13 (ig/dL, significantly decreased the
time spent exploring the center of the arena compared with controls shortly after exposure was terminated
(Wang et al.. 2016). However, Shvachiv et al. (2018) found no significant effect of developmental Pb
exposure on adult Wistar rats using the same measure, despite employing a longer exposure paradigm that
resulted in higher BLLs than Wang et al. (2016). Interestingly, Abazvan et al. (2014) reported OFT
findings suggestive of an anxiolytic effect of Pb exposure (i.e., increased central activity and increased
rearing) in male transgenic mice that were heterozygous for mDISCl (associated with increased risk for
psychiatric disorders including schizophrenia) but phenotypically normal.
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Six recent studies have evaluated the potential anxiogenic effects of Pb using the elevated plus
maze (EPM). The EPM is comprised of four arms—two closed and two open (i.e., with or without
walls)—and anxious behavior is indicated by an increase in the preference for the closed arms (or
inversely, decreased preference for the open arms). Despite not finding conclusive results indicative of
increased anxiety in the OFT, Shvachiv et al. (2018) found that Pb-treated Wistar rats (male and female
adults exposed consistently or intermittently since gestation) spent significantly less time in the open arms
of the EPM. This finding was corroborated in a subsequent study by the same research group, which
investigated lifetime Pb exposure in Wistar rats and found that the percent of time animals spent in the
open arms significantly decreased at 12, 20, and 28 weeks of age (Shvachiv et al.. 2020). Interestingly,
the greatest decrease in open arm presence was observed at 20 weeks. Abazvan et al. (2014) also
demonstrated an anxiety-like phenotype using the EPM in 6-month-old male and female mice following
lifetime exposure to Pb. Tartaglionc et al. (2020) found that exposure to Pb from gestation to weaning
significantly decreased entries into the open arms, decreased head dipping behavior and stretch-attend
postures in female Wistar rats at PND 60 (mean BLLs of 25 (.ig/dL): however, only the decreases in
stretch-attend postures were observed in males. One study, Neuwirth et al. (2019a). found no significant
behavioral differences in the EPM in adolescent Long-Evans rats following gestational and
developmental exposure in either dosing group (peak BLLs 3-11 (ig/dL for lower dose group and 9-18
(ig/dL for higher dose group).
Sobolewski et al. (2020) investigated the potential for transgenerational effects of Pb on this
endpoint by exposing female C57BL/6J mice (F0) prior to mating and during gestation, resulting in
offspring (Fl) with BLLs of 10-15 (ig/dL at PND 6-7. The developmentally exposed F1 generation was
paired with unexposed mice at PND 60 to produce the F2 generation, and the process was repeated to
produce the F3 generation which had no direct Pb exposure. F3 females spent significantly more time in
the open arms of the EPM. This effect could be further traced to descendants of the F1 sire line instead of
the Fl dam line. No significant effects were detected in F3 males.
The influence of Pb exposure on rodent behavior in the forced swim test (FST) and tail
suspension test (TST) has also been evaluated in recent studies. These tests are classically considered
models of emotional despair, with animals exhibiting both escape-directed behaviors and periods of
immobility (e.g., floating or hanging). Originally used to screen for antidepressant drugs, decreases in
immobility in the FST or TST following chemical exposure are interpreted as an antidepressant effect;
however, it was recently suggested that immobility is instead an adaptive response to the acute stress of
the FST or TST, and decreased immobility may be reflective of a maladaptive coping strategy or,
potentially, an anxiety-like phenotype (Anvan and Amir. 2018; Molendijk and de Kloet. 2015). Corv-
Slechtaet al. (2013) reported that C57BL/6 mice which had been exposed to Pb from gestation to
adulthood had significantly decreased immobile bouts in the FST compared with control animals. In
another recent study, postnatal exposure to Pb in male and female CD1 mice significantly increased their
time spent resisting in the TST (Duan et al.. 2017). These recent results indicate that, at least under some
experimental testing paradigms (producing mean BLLs as low as roughly 6 (.ig/dL). Pb exposure results in
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what has classically been considered an antidepressant effect but may be more aptly attributed to an
altered response to stress.
3.5.4.2.1 Summary
Studies in the previous ISA consistently supported increases in emotionality in rodents and
nonhuman primates following developmental Pb exposure that produced mean BLLs as low as 13 (ig/dL.
Recent studies largely support and expand on this conclusion. Consistent decreases in rodent exploratory
behaviors in the OFT and hole board test (e.g., rearing, sniffing, head dipping) were found in Pb-exposed
rodents with peak BLLs from 3 to greater than 30 (ig/dL, lower than previously demonstrated. An
anxiety-like phenotype was also demonstrated in the EPM by multiple studies, with only one study
reporting null effects. Sobolewski et al. (2020) also demonstrated potential sex-specific transgenerational
effects of Pb exposure on this endpoint. Inconsistent effects of Pb on thigmotactic behavior were reported
by a few studies, which was not an endpoint discussed in the previous ISA. Two studies demonstrated
decreased immobility in classical tests of depression-like behavior, suggestive of an antidepressant effect,
but the relevance of these tests to human depression is unclear. While limited studies reported null results,
they were not stronger with respect to design or methodology and did not significantly weaken the larger
body of evidence.
3.5.4.3 Relevant Issues for Interpreting the Evidence Base
3.5.4.3.1 Concentration-Response Function
Bayesian kernel machine regression (BKMR) and five-chemical linear regression models were
used to examine covariate adjusted associations between Pb exposure and CPRS Anxious-Shy T-score at
age 8 and BASC-second revision Anxiety T-score at age 15 Rokoff et al. (2022). BKMR analysis
indicated linear associations between Pb exposure and these outcomes, and no interactions between cord
Pb, Mn, and organochlorines.
3.5.4.3.2 Potentially At-Risk Populations
The 2013 Pb ISA did not describe populations of children potentially at higher risk of Pb-
associated internalizing behaviors. Recent epidemiologic studies presented sex-stratified results or
examined interactions between Pb exposure and other chemicals.
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Sex
Fruh et al. (2019) studied mother-child pairs participating in Project Viva, a longitudinal birth
cohort in eastern Massachusetts. This study found a stronger association of maternal BLL with the
emotional component of the SDQ measured in mid-childhood for girls compared with boys (|3 = 0.52
[0.18, 0.86] for girls v. (3 = 0.17 [95% CI: -0.17, 0.50] for boys). Note that higher scores on the emotional
problem scale indicate worse performance. In another study, Joo et al. (2018) found that late pregnancy
and cord BLL was associated with increasing internalizing behavior ratings on the CBCL in boys (|3 =
2.55 [95% CI: 0.22, 4.88] and |3 = 2.44 [95% CI: -0.74, 5.63], respectively), while postnatal (age 2 and 5)
BLL was associated with increasing internalizing behavior ratings on the CBCL in girls (|3 = 2.94 [95%
CI: 0.36, 5.52] and |3 = 5.65 [95% CI: 0.5, 10.8]). In a study that used Conners" rating scale to ascertain
internalizing behaviors, Rokoff et al. (2022) found the psychosomatic score was positively associated
with cord Pb and this association was stronger in boys than in girls (|3 = 2.08 [95% CI: 0.07, 4.10] versus
|3 = 0.48 [95% CI: -1.00, 1.97]). Of the experimental animal studies that evaluated both sexes, a small
number identified behavioral changes in Pb-exposed females while detecting minimal or no changes in
their male counterparts on the EPM, which could indicate that females are more sensitive to changes in
anxiety-like behavior after exposure to Pb (Sobolewski et al.. 2020; Tartaglionc et al.. 2020). Overall, no
consistent pattern was observed across the limited number of epidemiologic and toxicologic studies that
presented sex-stratified results. Each study used a different instrument to ascertain the outcomes.
Other Metals
A recent study examined the interaction effect between prenatal Pb exposure and other metals on
internalizing behavior scores on the BASC and the CPRS. BKMR analysis indicated no interactions
between cord blood Pb, Mn, and organochlorines that would indicate a deviation from additivity in a
study by Rokoff et al. (2022).
3.5.4.3.3 Lifestages
Epidemiologic studies consistently show that BLLs measured during various lifestages and time
periods, including the prenatal period, early childhood, and later childhood, and averaged over multiple
years, are associated with increases in internalizing behaviors. The identification of critical lifestages and
time periods of Pb exposure is complicated further by the fact that BLLs in older children, although
affected by recent exposure, are also influenced by Pb stored in bone due to rapid growth-related bone
turnover in children relative to adults. Thus, associations of neurodevelopmental effects with concurrent
BLL in children may reflect the effects of past and recent Pb exposures. Recent prospective studies add to
the evidence from the strongest studies in the 2013 Pb ISA that found associations with childhood
average blood and tooth Pb levels in children. These recent studies found associations between
internalizing behaviors and early childhood, maternal, and cord BLLs. Toxicological studies also provide
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support that the sensitive exposure window is not limited to a single phase of development. Rather,
effects of Pb exposure on anxiety or depression-like behavior in animals have been found following
gestational and postnatal exposure, exposure starting in adolescence, and lifetime exposure.
3.5.4.4 Summary and Causality Determination of Internalizing Behaviors in Children
The 2013 Pb ISA concluded that a causal relationship was likely to exist between Pb exposure in
children and internalizing behaviors based on the available evidence (U.S. EPA, 2013). Prospective
studies demonstrated associations between higher average blood (roughly 14 (ig/dL) or tooth Pb (i.e.,
reflective of prenatal or early postnatal Pb exposure depending on the tooth layer analyzed) levels and
higher parent and teacher ratings of internalizing behaviors, including withdrawn behavior and symptoms
of depression, fearfulness, and anxiety in children (aged 8-13). These associations were present after
adjustment for SES, birth outcomes, and parental education, but some uncertainty regarding potential
confounding by parental caregiving quality remained. Results from cross-sectional studies evaluating
lower concurrent BLLs (5 (ig/dL) were inconsistent. Increased emotionality in rodents and monkeys was
demonstrated at BLLs as low as 13 (ig/dL after exposure to Pb during development, and biological
plausibility was supported by findings of alterations in the HPA axis and dopaminergic and GABAergic
systems.
Several recent longitudinal epidemiologic studies with high to moderate participation rates relied
on an expanded array of instruments to assess internalizing behaviors (i.e., CBCL, SDQ, PRS, CTRS, and
BASC-2) compared with the studies in the 2013 Pb ISA. These studies observed associations with blood
Pb exposure (early childhood and prenatal BLLs <7 (.ig/dL). A limited number of studies evaluated child
sex (Rokoff et al., 2022; Fruh et al., 2019; Joo et al., 2018) as an effect modifier but were not consistent
with regard to sex-specific effects. The majority of analyses controlled for important potential
confounders including the quality of parental caregiving (Rokoff et al., 2022; Fruh et al., 2019), maternal
education, and SES (Rokoff et al., 2022; Fruh et al., 2019; Winter and Sampson, 2017; Liu et al., 2014b);
however, each potential confounder was not uniformly considered across studies. No association between
blood Pb and internalizing behaviors was observed among children with CKD or in prospective studies of
Pb concentration in blood (Sioen et al., 2013), teeth (Horton et al., 2018) or toenails (Doherty et al.,
2020), which reported relatively low participation rates. The limited number of studies that aimed to
distinguish types of internalizing behaviors indicated associations with the anxiety component (Rokoff et
al„ 2022; Rasnick et al„ 2021). Recent studies that found associations with prenatal or cord BLLs add to
the evidence. Uncertainty remains, however, regarding the exposure patterns associated BLLs in older
children and adults.
Recent experimental animal studies provide coherence with the previous findings that moderate
to high peak BLLs (12 to >30 (ig/dL) increase anxiety-like behaviors on the EPM, hole board test, and
OFT following Pb exposure during a single developmental window (including prenatal (Basha and
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Reddv. 2015). postnatal (Basha et al.. 2014). or adolescent periods (Wang et al.. 2016)) or throughout
development and beyond (Shvachiv et al.. 2020; Tartaglionc et al.. 2020; Shvachiv et al.. 2018; Abazvan
et al.. 2014). Overall, experimental animal studies provide more extensive support for anxiety-like
behaviors than for depression-like behaviors. However, two recent studies reported that Pb exposure
decreased immobility in classical tests of emotional despair following postnatal or lifetime Pb exposure
(Duan et al.. 2017; Corv-Slechta et al.. 2013). In addition to the well demonstrated effects at moderate to
high BLLs, two recent studies found altered behaviors in a nose poke task and FST following Pb
exposures resulting in low BLLs (3.2-10 (.ig/dL); moreover, one study was able to demonstrate exposure-
response relationships (i.e., higher BLLs were associated with greater behavioral changes) (Flores-
Montovaand Sobin. 2015).
Overall, the evidence is sufficient to conclude that there is likely to be a causal relationship
between Pb exposure and internalizing behaviors in children. This determination is based on
consistent evidence from both recent and past prospective epidemiologic studies, which demonstrate
positive associations between average blood Pb (prenatal, early childhood, lifetime) or childhood tooth Pb
levels (generally reflecting prenatal or early postnatal exposure) and multiple measures of internalizing
behaviors in children (aged 4-17) after adjustment for multiple confounding factors (e.g., SES, birth
outcomes, parental education). Recent toxicological studies provide further support for anxiety-like
behaviors following developmental and cumulative exposures that result in BLLs that are relevant to
humans. Despite these findings, some uncertainties have not been addressed in the epidemiologic
literature, including full consideration of certain confounding factors (e.g., parental caregiving quality)
and uncertainty regarding the exposure patterns associated with observed BLLs. Furthermore,
inconsistencies remain in the limited number of cross-sectional studies available in populations with
BLLs below 5 (ig/dL.
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Table 3-5 Summary of evidence for a likely to be causal relationship between Pb exposure and internalizing
behaviors in children
Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with
Effects0
Consistent results from
prospective epidemiologic
studies with relevant
exposures
Evidence from prospective studies for
higher ratings of internalizing behaviors in
children ages 8-13 yr in Boston and Port
Pirie cohorts in association with tooth or
lifetime average BLLs.
Section 4.3.4.1, (U.S. EPA, 2013)
Burns et al. (1999)
Bellinger et al. (1994b)
Blood Pb lifetime (to age 11-13 yr)
average mean: -14 |jg/dL
Tooth Pb (age 6 yr) mean: 3.4 |jg/g
Evidence from prospective studies for
higher rating of internalizing behaviors in
children 6-17 yr (cohorts in eastern MA,
Chicago, Cincinnati, and China) in
association with early childhood and
prenatal BLLs.
Winter and Sampson (2017)
Liu et al. (2014b)
Fruh et al. (2019)
Rokoff et al. (2022)
Early childhood <7 |jg/dL (median/mean)
Maternal and cord blood Pb, <2 |jg/dL
(median)
Associations also found in children aged
4-5 yr in former Yugoslavia in association
with lifetime average BLL
Wasserman et al. (2001)
Blood Pb lifetime (to age 4-5 yr) average
mean: 7.2 |jg/dL
Prospective studies had population-based
recruitment with moderate follow-up
participation. Participation not conditional
on tooth/BLLs and behavior
Inconsistent results in cross-sectional
studies with mean BLLs < 5
Section 4.3.4.1, (U.S. EPA, 2013)
Uncertainty regarding Epidemiologic associations found with Section 3.7, Table 3-10E
potential confounding adjustment for SES, birth outcomes,
parental education. Studies did not
uniformly adjust for parental caregiving
quality.
Uncertainty regarding the Uncertainty in regarding past exposure in
exposure patterns associated older children,
with observed BLLs.
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Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with
Effects0
And, supporting animal
evidence with relevant
exposures from multiple
studies
Gestational, lactational, and adolescent
exposures increasing anxiety-like
behaviors and altered stress coping
response.
Corv-Slechta et al. (2013)
Flores-Montova and Sobin (2015)
Shvachiv et al. (2020)
Peak BLLs: 3-27 pg/dL
BLL = blood lead level; Pb = lead; yr = year(s); SES = socioeconomic status.
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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3.5.5
Motor Function in Children
The evidence assessed in the 2013 Pb ISA is sufficient to conclude that a "causal relationship is
likely to exist" between Pb exposure and decrements in motor function in children. Evidence from
prospective studies of Cincinnati and Yugoslavia birth cohorts indicated associations of decrements in
fine and gross motor function with higher neonatal, concurrent, and lifetime average BLLs in children
aged 4.5-6 years and with higher earlier childhood (ages 0-5 years on average, age 78 months) BLLs in
children aged 15-17 years (Bhattacharya et al„ 2006; Ris et al., 2004; Bhattacharya et al., 1995; Dietrich
et al., 1993). The means for these blood Pb metrics ranged from 4.8 to 12 (ig/dL. These associations were
found with adjustment for several potential confounding factors, including SES, parental caregiving
quality, and child health with no indication of substantial selection bias. Evidence from cross-sectional
studies was less consistent, however (see Section 4.3.8 of (U.S. EPA. 2013)). The biological plausibility
for associations observed in children was supported by a study that found poorer balance in male mice
with relevant gestational to early postnatal (PND 10) Pb exposures. Overall, the strongest evidence was
from a small number of prospective cohort studies of children with limited support from studies in mice
with relevant exposures.
Measures of central tendency for Pb biomarker levels used in each study, along with other study-
specific details, including study population characteristics and select effect estimates, are highlighted in
Table 3-1 IE (Epidemiologic Studies) and Table 3-1 IT (Toxicological Studies). An overview of the recent
evidence is provided below. Overall, recent epidemiologic studies support findings from the 2013 Pb ISA
and a limited number of recent experimental animal studies provide coherence for their observations
demonstrating effects at relevant exposure concentrations.
3.5.5.1 Epidemiologic Studies of Motor Function
Evidence from prospective studies of Pb exposure and decrements in motor function in the 2013
Pb ISA indicated associations between higher neonatal, concurrent and lifetime average BLLs and motor
function decrements. Several recent epidemiologic studies examined the association between Pb exposure
and decrements in motor function in children. The findings generally support an association between Pb
exposure and decrements in motor function; however, they varied by the specific measure of motor
function as well as the timing of exposure measurement. Most studies were cohort studies and assessed
motor function using a comprehensive motor score, such as the Psychomotor Developmental Index (PDI)
score, from a version of the BSID (Jiang et al.. 2022; Kao et al.. 2021; Rygiel et al.. 2021; Shekhawat et
al.. 2021; Kim et al.. 2018b; Y Ortiz et al.. 2017; Paraiuli et al.. 2015b; Paraiuli et al.. 2015a; Liu et al..
2014c; Kim et al.. 2013c; Henn et al.. 2012). A few studies used a motor score from the Chinese version
of the GDS (Liu et al.. 2022a; Zhou et al.. 2017). The remaining studies assessed specific tasks, such as
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balance, manual dexterity, coordination, and fine motor speed (Taylor et al.. 2018; Boucher et al.. 2016;
Taylor et al.. 2015).
Studies using the Bayley scales to measure motor function in infants and toddlers (i.e., through
age 3) generally found associations between some Pb exposure metrics and decreased motor score. Kim et
al. (2013c). Kim et al. (2018b). Y Ortiz et al. (2017). Liu et al. (2014c). Rvgiel et al. (2021). and
Shekhawat et al. (2021) observed a decrease in motor score using maternal or cord BLLs, as well as other
blood Pb metrics, in several birth cohorts in multiple countries. Associations between BLLs and PDI are
presented in Figure 3-13.
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Study
Prospective Studies
TKim etal.2013
tKimetal.2018
tLiu etal. 2014
tRygiel et al. 2021
Location
3 Cities, S Korea
4 Cities, S Korea
Guangdong, China
Mexico City, Mexico
Claus Henn et al. 2012 Mexico City, Mexico
Blood Pb
Prenatal (late pregnancy)
Prenatal {early pregnancy)
Prenatal (late pregnancy)
Prenatal (early pregnancy)
Prenatal (late pregnancy)
Prenatal (late pregnancy)
Prenatal (late pregnancy)
Prenatal (late pregnancy)
Prenatal (cord)
Prenatal (T1)
Prenatal (T2)
Prenatal (T3)
Prenatal (T1)
Prenatal (T2)
Prenatal (T3)
Child (12 months)
Child (24 months)
Mean
(Mg/dl_)
Age at Outcome
(months)
1.4 (GM)
6
13 (GM)
6
1.4 (GM)
6
1.3 (GM)
6
1.4 (GM)
6
13 (GM)
6
2.7 (median)
13-24
NR
13-24
NR
13-24
5.63 (ret: 1.35)
36
5.27 (GM)
12
4.74 (GM)
12
4.98 (GM)
12
5.27 (GM)
24
4.74 (GM)
24
4.98 (GM)
24
5.1
12-36
5
12-36
Strata
Cd<1.47 pg/L
Cd>1.47 pg/L
Cd<1.51 pg/L
Cd >1.51 pg/L
Boys *
Girls *
Pb s3.92 vs. *1.89 pg/dL
—i 1 1 1
-15.00 -10.00 -5.00 0.00
Beta values (95% CI) per 1 ug/dL increase in blood Pb
Figure 3-13 Associations between biomarkers of Pb exposure and Bayley Score of Infant Development
Psychomotor Developmental Index.
Note; Effect estimates are standardized to a 1 pg/dL increase in blood Pb or a 10 (jg/'g increase in bone Pb. If the Pb biomarker is log-transformed, effect estimates are standardized to
the specified unit increase for the 10th -90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval. Categorical effect
estimates are not standardized. Associations that could not be standardized are not included on the plot.
tStudies published since the 2013 Integrated Science Assessment for Lead.
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Kim et al. (2013c) found that PDI score at 6 months of age decreased with increasing BLLs
measured in the third trimester (median = 39th week) (|3 = -1.38 [95% CI: -3.31, 0.55] per 1 (ig/dL
increase in BLL) in the Korean MOCEH study. Another Korean study using the CHECK cohort (Kim et
al.. 2018b) also observed decrements in PDI among 13-24 month old infants in association with perinatal
maternal BLLs (|3 per l-(ig/dL blood Pb = -15.45 [95% CI: -30.12, -0.79]). Sex-stratified results were
slightly negative but not significant. In China, Liu et al. (2014c) observed an association between
increasing prenatal (umbilical cord blood) Pb levels and worse PDI score at 36 months of age. Compared
with low prenatal Pb (<1.89 (.ig/dL). children exposed to high prenatal Pb (>3.92 (ig/dL) were more likely
to have a lower PDI score (|3 = -1.30 [95% CI: -1.57, -1.03]), after adjusting for potential confounders.
Several studies in Mexico also examined the association of BLL with PDI assessed in infants.
Rvgiel et al. (2021) found a small negative association between prenatal (trimester-specific) BLLs and
PDI scores at 12 months in the ELEMENT Project study (|3 per 1 (ig/dL increase in 1st trimester Pb =
-0.24 [95% CI: -0.95, 0.48]; |3 per 1 (ig/dL increase in 2nd trimester Pb = -0.38 [95% CI: -1.10, 0.35]; |3
per 1 (ig/dL increase in 3rd trimester Pb = -0.33 [95% CI: -1.06, 0.40]). At 24 months, the negative
association persisted but with a smaller magnitude of effect. Rvgiel et al. (2021) also examined whether
DNA methylation mediated the association and found that DNA methylation of cgl8515027 located
within glucosaminyl (N-acetyl) transferase 1 (GCNT1) had a suppressive (positive indirect) effect on the
inverse relationship between second trimester BLLs (ln-transformed) and PDI scores at 12 months (|3indirect
= 1.25 (95% CI: -0.11, 3.32]). In the Programming Research in Obesity, Growth, Environment and
Social Stressors (PROGRESS) birth cohort in Mexico, Y Ortiz et al. (2017) found a negative association
between motor score at 24 months of age and log-transformed BLLs measured during the third trimester
(|3 = -11.01 [95% CI: -17.55, -4.48]), but not for BLLs measured during the second trimester (|3 = 1.97
[95% CI: -2.46, 6.40]). In a study using childhood BLLs, Henn et al. (2012) found a small negative
association with repeated measures of PDI scores in another cohort of children in Mexico. From adjusted
mixed-effects models with repeated measures of PDI scores at 12, 18, 24, 30, and 36 months, there was a
negative association between 12-month blood Pb and PDI scores (|3 per 1 -(.ig/dL blood Pb = -0.27 [95%
CI: -0.56, 0.02]). Similarly, from adjusted mixed-effects models with repeated measures of PDI scores at
24, 30, and 36 months, there was a negative association between 24-month blood Pb and PDI scores (|3
per 1-ng/dL blood Pb = -0.18 [95% CI: -0.53, 0.17]).
Several studies are not pictured in Figure 3-13. Shekhawat et al. (2021) found that children with
cord blood Pb concentrations of 5-10 (ig/dL had reduced gross motor skills on the BSID at an average
age of 6.5 months (|3 = -0.29 [95% CI: -5.00, 0.11]) for each 1 (ig/dL increase in cord BLL. Additionally,
in a birth cohort of mother-child pairs recruited from Bharatpur General Hospital in Nepal, Paraiuli et al.
(2015a) and Paraiuli et al. (2015b) assessed the association of cord BLLs with PDI at 24 months old and
36 months of age, respectively. Negative but non-significant associations were observed between log-
transformed cord BLLs and 24-month PDI (|3 = -4.83 [95% CI: -16.53, 6.86]) nor 36-month PDI (|3 =
-2.56 [95% CI: -9.71, 4.59]). Notably, two additional studies that used biomarkers other than blood did
not find associations, i.e., Jiang et al. (2022) measured Pb in meconium (at birth) and in hair and
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fingernails (at 3 years of age) in Taiwan and did not find an association with any motor score (total, fine
motor, or gross motor) at 3 years of age. Another study in Taiwan (Kao et al.. 2021) that used hair and
fingernail biomarkers of Pb concentrations similarly did not report significant associations with motor
development among infants less than 3 years old.
Several other instruments were used to assess motor function in infants and toddlers. Paraiuli et
al. (2013) measured Pb, As, and Zn levels in cord blood and used the third edition of the Brazelton
Neonatal Behavioral Assessment Scales (NBAS III) to assess neurodevelopment in one-day-old newborns
in Chitwan, Nepal. The NBAS III contains 27 behavioral and 18 reflex items and is used for infants up to
2 months old. The multivariate model was adjusted for parity, family income, mother's age, education,
BMI, birth weight, gestational age, and age in hours at NBAS assessment. The NBAS motor cluster score
was inversely associated with the log-transformed cord BLLs (|3 = -2.15 [95% CI: -4.27, -0.03]). Liu et
al. (2014d) used the Neonatal Behavioral Neurological Assessment (NBNA), which is based on the
NBAS and has five clusters of behavior: passive tone, active tone, primary reflexes, and general
assessment. Newborns in this study were assessed at 3 days old, and the NBNA has been validated among
Chinese newborns between 2 and 28 days old. Associations between maternal BLL in the first trimester
and the NBNA scores were observed (|3 = -4.86 [95% CI: -8.83, -0.89] per unit of log-transformed Pb).
Less precise associations of second trimester, third trimester, and cord BLLs with decreased motor
function were also observed. Among toddlers (2-3 years old) Zhou et al. (2017) and Liu et al. (2022a)
both used the Chinese version of the GDS to calculate a motor score. For every loglO (|ig/dL) increase in
maternal blood Pb (measured at 28-36 weeks of gestation), Zhou et al. (2017) observed a positive
association for gross motor development ([3 = 3.31 [95% CI: -6.11, 12.73] per log-10 transformed unit of
BLL) as well as fine motor development (|3 = 0.49 [95% CI: -11.27, 12.24] per log-10 transformed unit
of BLL); however, the effect estimates were extremely imprecise. On the other hand, for each In (|ig/L)
increase in maternal Pb, Liu et al. (2022a) observed a negative association for gross motor development
(|3 = -2.32 [95% CI: -3.61, -1.03] per ln-transformed unit of BLL). Furthermore, Nvanza et al. (2021)
did not find associations between high Pb exposure and fine or gross motor impairment assessed by the
MDAT.
Several additional studies were conducted using assessment instruments that measure children's
(7 years or older) ability to perform certain tasks. In the ALSPAC, Taylor et al. (2015) conducted a heel-
to-toe test in children at age 7 years, beam walking test (to measure dynamic balance) at age 10 years, and
balancing test with eyes closed (to measure static balance) also at age 10 years. Pb levels measured in
maternal blood (<18 weeks of gestation) and Pb levels measured in child blood (30 months old) were not
associated with any measure of motor function in this study (Taylor et al.. 2015). In another analysis of
ALSPAC data, Taylor et al. (2018) examined the association between first trimester BLLs and different
measures of coordination. Compared with prenatal blood Pb <5 (ig/dL, children exposed to higher levels
(>5 (ig/dL) of prenatal Pb were more likely to fail the tests of manual dexterity (threading lace, peg board
using preferred hand, and peg board using non-preferred hand). When comparing the highest blood Pb
quartile to the lowest blood Pb quartile, the only association remaining was for failing the peg board using
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the preferred hand (OR for quartile 4 versus quartile 1 = 1.23 [95% CI: 0.92, 1.66]). Prenatal Pb exposure
was not associated with tests of balance and the results were inconsistent for ball skills (inverse
association for >5 (ig/dL versus <5 (ig/dL; positive association for quartile 4 versus quartile 1). In the
Nunavik Child Development Study in Canada, Boucher et al. (2016) measured manual dexterity, fine
motor speed, and visuomotor integration in children (ages 8.5-13.3 years). BLLs (log-transformed)
measured at birth (cord blood) and at age 11 years were negatively associated with manual dexterity (|3
for cord blood Pb = -0.08 [p > 0.10]; (3 for child blood Pb = -0.17 [95% CI: -0.34, 0.00]) and fine motor
speed (|3 for cord blood Pb = -0.19 [95% CI: -0.33, -0.05]; |3 for child blood Pb = -0.21 [95% CI: -0.37,
-0.05]). For visuomotor integration, there was no association with cord blood Pb (|3 for cord blood Pb =
-0.01 [p > 0.10]) and apositive association with child blood Pb (|3 for child blood Pb = 0.10 [p > 0.10]).
The magnitude of effect was greater for child BLLs. Nozadi et al. (2021) collected blood samples from
pregnant mothers at the 36-week visit or at the time of delivery and administered the ASQ:I at 10-13
months of age to evaluate communication, gross motor, fine motor, problem-solving, and personal-social
development. A 1 -(.ig/dL increase in prenatal blood Pb was associated with a decrease in fine motor (|3 =
-0.63 [95% CI: -1.19, -0.08]) scores. Palaniappan et al. (2011) observed decrements ofWRAVMA
scores in association with l-(ig/dL increase in concurrent BLLs (Drawing: |3 = -0.29 [95% CI: -0.51,
-0.07]; Matching: |3 = -0.14 [95% CI: -0.31, 0.02]; Pegboard: |3 = -0.19 [95% CI: -0.38, 0.01];
Composite: |3 = -0.26 [95% CI: -0.45, -0.07]).
3.5.5.1.1 Summary
Evidence from prospective studies of Cincinnati and Yugoslavia birth cohorts indicated
associations of decrements in fine and gross motor function with higher neonatal, concurrent, and lifetime
average BLLs in young children with higher earlier childhood BLLs. Several recent birth cohort studies
observed lower scores on the Bayley PDI in association with maternal Pb exposure (no clear pattern by
trimester of pregnancy), cord BLL, and postnatal concurrent blood Pb (Rygiel et al.. 2021; Y Ortiz et al..
2017; Liu et al.. 2014c; Kim et al.. 2013c; Henn et al.. 2012). Pb-associated decrements in motor function
were observed in neonates (Liu et al.. 2014d; Paraiuli et al.. 2013) and in some but not all studies of
toddlers that assessed motor function using GDS (Liu et al.. 2022a; Zhou et al.. 2017) or children's
(greater than 7 years old) abilities to perform certain tasks indicative of gross motor function (i.e.,
balance) (Taylor et al.. 2015). although associations with fine motor function were observed (Taylor et al..
2018; Boucher et al.. 2016).
3.5.5.2 Toxicological Studies of Motor Function
As described above and in previous reviews (U.S. EPA. 2013. 2006). epidemiologic studies
provide evidence of associations between Pb exposures and fine and gross motor decrements, mainly in
children. Evaluating performance in neurobehavioral toxicological studies with Pb exposure in rodents
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can substantiate observed Pb exposure effects on motor function seen in humans. In past assessments,
evidence in animal toxicological studies has been limited due to a lack of investigations with relevant Pb
exposures. The purpose of this section is to update the collection of evidence available concerning Pb
exposure-induced effects on motor function in animal models. Studies that examined various indices of
locomotor activity are evaluated above in the Toxicological Studies of Hyperactivity section (Section
3.5.2.3.2).
Previous IS As (U.S. EPA, 2013, 2006) highlighted rotarod and air righting reflex experiments
with rodents to discuss the effects of Pb on development of motor coordination and balance. Typical
rotarod tests compare the latency to fall for subjects placed on a rotating rod. Falling off more quickly
indicates decreased coordination and/or balance. There are two rotarod studies discussed in previous U.S.
EPA reviews that describe effects of developmental Pb exposure on rotarod performance that resulted in
relevant BLLs less than 30 (ig/dL. Interestingly, Moreira et al. (2001) saw no effect of Pb exposure, from
the beginning of gestation through lactation, on Wistar rat rotarod performance at PND 70 with PND 23
mean BLLs of 21 (ig/dL. In contrast, Leasure et al. (2008) observed substandard performance in
pregestational through lactation Pb-exposed male, but not female, mice with peak BLLs of less than 10
(ig/dL. Since Leasure et al. (2008), no other PECOS-relevant studies have assessed rotarod performance
in rodents exposed to Pb throughout the entire developmental period. Two recent rotarod studies with
mice, by Flores-Montoya and Sobin (2015) and Zou et al. (2015), showed no decrements in performance
in rotarod tests after postnatal-only exposure to Pb in drinking water for PND 0-28 and 37-58 for
respective studies.
While the outcomes of these two latest rotarod studies were mostly negative, additional
investigations evaluating the effects of developmental Pb exposure on coordination and balance in
neonatal rats using surface righting reflex, negative geotaxis reflex, and ascending wire mesh tests yielded
mixed results. Surface righting reflex tests are run by placing pups in a supine position and then recording
the time it takes to flip onto their feet. Slower times to flip indicate postural imbalances. For negative
geotaxis reflex, or slant-board tests, pups are placed on a slanted board and the time it takes for the pup to
face upward is recorded. Slower times to turn upward indicate that the vestibular response to gravity cues
or motor coordination required for turning are underdeveloped. Success in ascending wire mesh tests also
requires coordination, as the animals are required to climb to the top of a mesh out of a water bath in a
predetermined period. In a study comparing the developmental effects on male Wistar rats with
pregestational, gestational, or lactational Pb exposure, pups exposed during gestation achieved negative
geotaxis significantly faster than unexposed counterparts when tested on PND 8, 10, and 12 (Rao Barkur
and Bairy, 2016). In contrast, in the same study, Rao Barkur and Bairy (2016) observed no difference in
negative geotaxis times between control, pregestation alone, and lactation alone Pb-exposed pups. No
effects on surface righting reflex on PND 3 through 5 were observed for pups belonging to the previously
mentioned exposure groups. The day of achievement in ascending wire mesh tests (PND 14-18) was
delayed for animals in both gestation and lactation Pb-exposed groups but not for those in the
pregestational group (Rao Barkur and Bairy, 2016). Betharia and Maher (2012) exposed pregnant
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Sprague Dawley rats to Pb (II) acetate trihydrate via drinking water from the beginning of gestation
through lactation and until weaning. Development of the surface righting reflex of control and exposed
offspring was tested from PND 1 to 10. Slower righting times were observed for Pb-exposed offspring on
PND 1; however, from PND 2 through 10, there were no differences between Pb-exposed and control
groups. Basha and Reddv (2015) observed a significant increase in righting time in righting reflex tests
done on PND 6 and 7 and an increase in latency to turn in negative geotaxis tests for male Wistar rats
tested on PND 8, 9, and 10 after in utero exposure to Pb. Tartaglione et al. (2020) saw no decrements in
righting reflex time or negative geotaxis achievement on PND 4, 7, 10, and 12 from Pb exposure in the
offspring of dams exposed to Pb from 1 month pre-mating to offspring weaning.
Additional motor function experiments with early postnatal weaning in Wistar rats were carried
out in studies with developmental Pb exposures. Tartaglione et al. (2020) recorded on PND 4, 7, 10, and
12 the duration of neonatal motor patterns of rat pups from dams exposed to Pb before mating until
offspring weaning. On PND 10, Pb-exposed pups spent less time in locomotion compared with controls,
in favor of head rising and wall climbing movements, demonstrating a stereotyped/preservative profile.
Basha and Reddv (2015) observed a prenatal Pb-induced strength deficit when rats were subject to
fore limb hang tests on PND 13, 14, 15, and 16 but not on day 12. This indicated Pb-induced
underdevelopment of fine motor ability. Rao Barkur and Bairv (2016) tested rats on PND 6, 8, 10, and 12
for Pb-induced effects on swimming development. They observed no difference in swimming body angle
or limb movements for ISA-relevant pregestation, gestation, or lactation-exposed groups compared with
control. These novel studies warrant further investigation into the effects of Pb exposure at different
concentrations and stages of development on neonatal movement patterns and forelimb hang tests.
3.5.5.2.1 Summary
The evidence supporting the link between developmental Pb exposure and deficits in motor
function in animal models has expanded on account of recent studies utilizing Pb-exposed rodents with
mean BLLs <30 (ig/dL. These new studies illustrate the effects of Pb exposure on both gross and fine
motor development in novel paradigms. In addition to the effect on rotarod performance (Leasure et al..
2008) described in the previous ISA, developmental Pb-induced decrements in righting reflex, negative
geotaxis reflex (Basha and Reddv. 2015). ascending wire mesh (Rao Barkur and Bairv. 2016). and
forelimb hang tests (Basha and Reddv. 2015) were observed. Interestingly, gestational Pb exposure was
present among each type of study that yielded decrements in these measurements of motor function;
therefore, it may be a more sensitive window compared with lactation or postnatal exposures. In terms of
design or methodology, studies that found weak or null relationships were not stronger and did not
weaken the overall body of corroborating data. Key aspects such as exposure levels and timing, ages of
animals at testing, and slant-board angles were variable between the few relevant studies. Altogether, the
results from these recent studies support the conclusions from the previous ISA. However, due to the
limited number of reproduced experiments, these recent studies do not enhance the consistency of the
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evidence. In addition to the fine motor, motor reflex, and coordination, and balance studies described in
this section, effects of developmental Pb exposure on locomotor activity are evaluated separately in the
Toxicological Studies of Hyperactivity section (Section 3.5.2.3.2) above. Briefly, due to heterogeneity in
study design, the evidence for effects of developmental Pb on locomotor activity is mixed; however, a set
of four independent studies with analogous conditions showed hyperactivity in rodents when tested within
a PND 14 to 23 window after lactational Pb exposure (Duan et al.. 2017; De Marco et al.. 2005; Moreira
et al.. 2001; Rodrigues et al.. 1996).
3.5.5.3 Relevant Issues for Interpreting the Evidence Base
3.5.5.3.1 Potentially At-Risk Populations
Sex
A limited number of toxicological studies have reported sex differences in Pb-related effects on
motor function. Among studies in the 2013 Pb ISA, sex-specific differences in mice were observed for
gross motor skills, with balance and coordination most affected among males at the lowest Pb exposures
(Leasure et al.. 2008).
Recent epidemiologic studies that evaluated sex as a potential modifier of the association between
Pb exposure and motor function add to the evidence (Liu et al.. 2022a; Y Ortiz et al.. 2017). Y Ortiz et al.
(2017) found that the observed association between maternal blood Pb during the third trimester and
lower PDI scores was not different between boys and girls. Liu et al. (2022a) found that the association of
maternal blood Pb exposure with gross motor development quotient on the GDS was modified by sex
(-3.43 [95% CI: -6.16,-0.69] in boys and-1.18 [95% CI: -2.81, 0.44] in girls per ln-transformed unit).
Maternal Self-esteem
Maternal self-esteem has been shown to modify associations between BLLs and health effects in
children. In one study, high maternal self-esteem appeared to attenuate the negative effects of the child's
increased BLLs on PDI scores (Surkan et al.. 2008). In this study, larger decreases in PDI scores were
associated with increased BLLs among children whose mothers were in the lower quartiles of self-esteem
(Surkan et al.. 2008). Maternal self-esteem was not evaluated as an effect modifier in recent studies of Pb
exposure and motor function among children.
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Maternal Stress
In a recent epidemiologic study, Y Ortiz et al. (2017) found that the observed association of
maternal blood Pb during the third trimester with lower PDI scores differed depending on maternal stress.
Contrary to expectations, higher PDI scores were observed with higher maternal stress.
3.5.5.3.2 Lifestages
Multiple lifestages during childhood are implicated in the effects of Pb exposure on motor
function in children. Analyses of children enrolled in the Cincinnati cohort at age 6 years indicated
associations of concurrent, lifetime average, and neonatal Pb exposure with poorer upper limb dexterity
and fine motor composite score. Studies conducted in the Cincinnati cohort found that prenatal or
neonatal BLLs were not consistently associated with motor function decrements at ages 4-10 years
(Bhattacharva et al.. 1995; Dietrich et al.. 1993). Several recent birth cohort studies support findings from
the 2013 Pb ISA with observations of lower scores on the Bayley PDI in association with maternal Pb
exposure (no clear pattern by trimester of pregnancy), cord BLL, and postnatal concurrent blood Pb
(Rvgiel et al.. 2021; Y Ortiz et al.. 2017; Liu et al.. 2014c; Kim et al.. 2013c; Henn et al.. 2012). Animal
toxicological studies mentioned above and in previous ISAs indicate the potential for delays in gross
motor development with gestational and/or early postnatal Pb exposure (Rao Barkur and Bairv. 2016;
Basha and Reddv. 2015; Leasure et al.. 2008) and for fine motor decrements with gestational Pb exposure
(Basha and Reddv. 2015). Apart from the study by Leasure et al. (2008). which tested balance in adults,
these studies measured and found diminished motor development in early postnatal rodents (Rao Barkur
and Bairv. 2016; Basha and Reddv. 2015).
3.5.5.4 Summary and Causality Determination: Motor Function in Children
The evidence assessed in the 2013 Pb ISA is sufficient to conclude that a "causal relationship is
likely to exist" between Pb exposure and decrements in motor function in children. Key evidence came
from prospective analyses of the CLS and Yugoslavia cohorts demonstrating associations of BLLs with
poorer motor function with consideration of potential confounders including SES, parental caregiving
quality and education, smoking birth outcomes, sex, and child health. Among children that participated in
the Cincinnati cohort, higher earlier childhood BLLs (age 0-5 year average [median: 11.7 |ig/dL| or age
78 month) were associated with poorer fine (i.e., grooved pegboard and finger tapping) (Ris et al.. 2004)
and gross motor function (i.e., postural balance) (Bhattacharva et al.. 2006) assessed in adolescence (ages
12, 15-17 years). In addition, assessments of children enrolled in the Cincinnati cohort at age 6 years
indicated associations of concurrent (mean: 10.1 (.ig/dL). lifetime average (mean: 12.3 (.ig/dL). and
neonatal (mean: 4.8 (ig/dL) but not prenatal maternal (mean: 8.4 (ig/dL) BLLs with poorer upper limb
dexterity, fine motor composite score (Dietrich et al.. 1993). and poorer postural balance (Bhattacharva et
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al.. 1995). Wasserman et al. (2000) also examined the association of Pb exposure with motor function. In
this prospective analysis of the Yugoslavian cohort, an association of lifetime average BLL (exact levels
not reported) with decrements in fine but not gross motor function at age 4.5 years was observed
(Wasserman et al.. 2000). Evidence from cross-sectional studies for associations between motor function
and concurrent BLL was mixed in populations with mean BLLs of 2-5 (ig/dL (Min et al.. 2007; Surkan et
al.. 2007; Despres et al.. 2005). Recent epidemiologic and toxicologic studies generally support findings
from the 2013 Pb ISA. The key evidence, as it relates to the causal framework, is summarized in Table
3-6.
Several recent birth cohort studies report lower scores on the Bayley PDI in association with
maternal Pb exposure (no clear pattern by trimester of pregnancy), cord BLL, and postnatal concurrent
blood Pb (Rygiel et al.. 2021; Y Ortiz et al.. 2017; Liu et al.. 2014c; Kim et al.. 2013c; Henn et al.. 2012).
Pb-associated decrements in motor function were also observed in neonates (Liu et al.. 2014d; Paraiuli et
al.. 2013). A limited number of studies of children greater than 7 years old were conducted. Taylor et al.
(2015) did not report associations with certain tasks indicative of gross motor function (i.e., balance),
although associations with decreased fine motor function were observed (Taylor et al.. 2018; Boucher et
al.. 2016).
Recent toxicological studies provide limited biological plausibility by showing effects on motor
function in rodent models from developmental Pb exposure resulting in BLLs <30 (ig/dL within one order
of magnitude of recent concentrations observed in humans. Epidemiologic evidence of developmental Pb-
induced impairment of balance and coordination is supported by observations of poorer rotarod
performance in male mice exposed to Pb during gestation (Leasure et al.. 2008). In addition, evidence
from epidemiologic studies indicating Pb-induced delayed gross motor development in children is
reinforced by toxicological studies that display slower times to achievement by postnatal rats
gestationally exposed to Pb in surface righting reflex (gestational Pb), negative geotaxis reflex
(gestational Pb) (Basha and Reddv. 2015). and ascending wire mesh tests (gestational Pb; lactational Pb)
(Rao Barkur and Bairv. 2016). Epidemiologic studies revealing Pb-induced decrements in children's fine
motor skills are supported by the observed grip strength deficits for gestational Pb-exposed early postnatal
rats in forelimb hang tests (Basha and Reddv. 2015). Additional studies on Pb-induced changes on several
neurochemical endpoints that factor into impaired motor function have been reported and are described in
Section 3.3
Overall, the evidence is sufficient to conclude that there is likely to be a causal relationship
between Pb exposure and motor function in children. This determination is based on consistent
evidence from prospective epidemiologic studies, which demonstrate an association between higher
childhood BLLs (neonatal, earlier childhood, concurrent and lifetime average) and poorer fine and gross
motor function in children (aged 4.5-17) with adjustment for maternal IQ, parental education, SES, and
HOME score. Additional prospective studies have also demonstrated consistent evidence in infants and
toddlers using the Bayley PDI, but evidence supporting neonatal effects is more limited. Epidemiologic
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evidence is supported by limited experimental animal studies that demonstrate impairments in balance,
coordination, and grip strength, as well as delayed reflex development. There is some remaining
uncertainty arising from cross-sectional studies using concurrent BLLs that have reported mixed results.
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Table 3-6 Summary of evidence indicating a likely to be causal relationship between Pb exposure and motor
function in children
Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with
Effects0
Consistent findings from a few
prospective epidemiologic studies
with relevant BLLs
Evidence from prospective studies for fine and gross
motor function decrements in children ages 4.5-17
yr in Cincinnati, Yugoslavia in association with
neonatal, earlier childhood, concurrent, lifetime avg
BLLs.
High follow-up participation, no selective attrition in
Cincinnati cohort, higher loss-to-follow-up in
Yugoslavia cohort with lower maternal IQ, HOME.
Both studies adjusted for maternal IQ, parental
education, SES, HOME score
Studies used various, widely used tests to assess
outcomes.
Ris et al. (2004)
Dietrich et al. (1993)
Bhattacharya et al.
(1995)
Wasserman et al.
(2000)
Section 4.3.7, (U.S.
EPA, 2013)
Blood Pb means
Cincinnati: neonatal (10 day) 4.8 |jg/dL,
concurrent (age 6 yr) 11.6 |jg/dL, lifetime
(to age 15-17 yr) avg 12.3 |jg/dL, age 0-5
yravg 11.7 |jg/dL
Former Yugoslavia: NR
Mixed evidence for lower (concurrent) BLLs from Section 4.3.7,
cross-sectional studies that considered several EPA, 2013)
potential confounding factors.
(U.S.
Consistent findings from prospective
studies of infants and toddlers
Lower scores on the Bayley PDI in association with
maternal Pb exposure (no clear pattern by trimester
of pregnancy), cord BLL and postnatal concurrent
blood Pb
Kim et al. (2013c)
Y Ortiz etal. (2017)
Liu et al. (2014c)
Rvaiel et al. (2021)
Henn etal. (2012)
Limited evidence in neonates
Pb-associated decrements in motor function
reflexes) observed
e.g., Paraiuli et al. (2013)
Liu etal. (2014d)
Limited experimental animal
evidence at relevant exposures
Deficient gross motor coordination and balance in
rodents with developmental Pb exposure (less time
on rotarod, slower righting and negative geotaxis
reflexes, delayed day of achievement for ascending
wire mesh test)
Leasure et al. (2008)
Basha and Reddv
(2015)
Rao Barkur and Bairv
(2016)
Blood Pb: -10 |jg/dL in mice after
pregestational through lactation exposure,
5-11 |jg/dL in rats after gestational
exposure, 27 |jg/dL in rats after lactational
exposure
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Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with
Effects0
Fine motor (grip strength) deficits in early postnatal
rats with gestational Pb exposure
Basha and Reddv
(2015)
Blood Pb: 11.2 |jg/dL after gestational
exposure
Limited experimental animal
evidence at relevant exposures
provide coherence for epidemiologic
observations of effect modification
by sex
Poorer balance (fell off rotarod more quickly) in adult
male but not female mice with pregestational
through lactation dietary Pb exposure
Leasure et al. (2008)
Blood Pb: -10 |jg/dL in mice after
pregestational through lactation exposure
Biological plausibility demonstrated
Pathways involving oxidative stress, inflammation
and Ca2+ signaling result in impaired neuron
development, synaptic changes, and
neurotransmitter changes.
Recent studies support and extend findings related
to overt nervous system effects
U.S. EPA (2013)
Section 3.3
Section 3.4.2
avg = average; BLL = blood lead level; Ca2+ = calcium ion; HOME = Health Outcomes and Measures of the Environment; IQ = intelligence quotient; NR = not reported; Pb = lead;
PDI = Psychomotor Developmental Index; SES = socioeconomic status; yr = year(s).
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 20151.
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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3.5.6
Sensory Organ Function in Children
The 2013 Pb ISA included separate causality conclusions for auditory and visual function. This
ISA combines these categories and makes one causality determination for Sensory Organ Function
because there are relatively few studies within this outcome grouping.
3.5.6.1 Auditory Function in Children
The evidence assessed in the 2013 Pb ISA was sufficient to conclude that "a causal relationship is
likely to exist" between Pb exposure and decrements in auditory function in children (U.S. EPA, 2013).
Evidence from a prospective study (Dietrich et al.. 1992) and small number of cross-sectional studies of
U.S. children, including NHANES and Hispanic Health and Nutrition Examination Survey (HHANES)
analyses (Schwartz and Otto, 1991, 1987) indicated associations of higher BLLs with increases in hearing
thresholds as well as decreases in auditory processing or auditory evoked potentials, with adjustment for
potential confounding by SES in most studies and by child health and nutritional factors in some studies.
The high participation rates in a prospective birth cohort study (Dietrich et al., 1992) reduced the
likelihood of biased participation by children with higher BLLs. Across studies, associations were found
with BLLs measured at various time periods, including prenatal maternal, neonatal (10 days, mean 4.8
(ig/dL), lifetime average (to age 5 years), and concurrent (ages 4-19 years) BLLs (median 8 (.ig/dL).
Evidence for Pb-associated increases in hearing thresholds or latencies of auditory evoked potentials was
also found in adult monkeys with lifetime dietary Pb exposure. However, these effects in adult animals
were demonstrated at higher peak or concurrent BLLs (i.e., 33-150 (ig/dL) than those relevant to this
ISA; thus, the biological plausibility for epidemiologic observations was unclear.
In the current ISA, several recent cross-sectional studies support the conclusion in the 2013 Pb
ISA regarding the association of Pb exposure with hearing loss; however, results were inconsistent for
other audiometric parameters. Recent toxicological studies provide additional evidence for hearing loss
and auditory processing deficits in rodents at relevant BLLs. Measures of central tendency for Pb
biomarker levels used in each study, along with other study-specific details, including study population
characteristics and select effect estimates, are highlighted in Table 3-12E (Epidemiologic Studies) and
Table 3-16T (Toxicological Studies). An overview of the recent evidence is provided below.
3.5.6.1.1 Epidemiologic Studies of Auditory Function
Several recent epidemiologic studies examined the association between Pb exposure and
decrements in auditory function in children. The findings generally support a positive association between
Pb exposure and hearing loss. For other audiometric parameters, however, the results were inconsistent.
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Most studies of auditory function were cross-sectional. In a meta-analysis of studies from Iran, Korea,
China, and the United States, Yin et al. (2021) observed a positive association between Pb exposure and
hearing loss indicated by pure-tone average (PTA) >25 dB in children and adolescents (3-19 years)
(combined OR per unit increase in Pb = 1.53 [95% CI: 1.24, 1.87]). The pooled OR was based on only
two studies (Xu et al.. 2020; Choi and Park. 2017). Xu et al. (2020) conducted a case-control analysis of
preschool-aged children (3 to 7 years of age) who resided in an area contaminated with Pb and Cd and in
an uncontaminated reference area. This study found associations of exposures with hearing loss,
potentially affected by epigenetic changes. The OR for Pb-associated hearing loss in both ears was 1.40
(95% CI: 1.06, 1.84 per unit change log-transformed BLL) after adjustment for characteristics of the
child, parental education, SES, and noise exposure. Choi and Park (2017) measured speech- and high-
frequency hearing loss in adolescents (12-19 years) and adults (20-87 years) in the Korea National
Health and Nutrition Examination Survey (KNHANES). Hearing loss was defined as PTA >15 dB in
adolescents. For each doubling of blood Pb, there was a positive association with speech-frequency
hearing loss (OR =1.2 [95% CI: 0.48, 3.05]) and high-frequency hearing loss (>25 dB) (OR = 1.26 [95%
CI: 0.73, 2.16]) among adolescents.
In addition to the aforementioned studies, among adolescent NHANES participants (ages 12 to
19 years), Shargorodskv et al. (2011) found a positive association between blood Pb and hearing loss.
Hearing loss was defined as low or high-frequency PTA >15 dB in either ear. Compared with study
participants with low BLLs (<1 (ig/dL), those with the highest level (>2 (ig/dL) were more likely to have
any hearing loss (OR = 1.95 [95% CI: 1.24-3.07]), particularly high-frequency hearing loss (OR = 2.22
[95% CI: 1.39-3.56]). The direction of effect for low-frequency hearing loss was the same but at a
smaller magnitude (OR= 1.13 [95% CI: 0.61-2.07]). Among younger children (3-7 years with median
BLL <5 (ig/dL) in China, a positive association between blood Pb and hearing loss was also observed
(Liu et al.. 2018c). For each (ig/dL increase in blood Pb, the odds of any hearing loss increased by 1.24
times (OR= 1.24 [95% CI: 1.03, 1.49]). This association was not as evident for high-frequency hearing
loss (OR= 1.08 [95% CI: 0.84, 1.38]) and low-frequency hearing loss (OR= 1.02 [95% CI: 0.87, 1.19]).
Auditory function in children was also measured according to the auditory brainstem response
(ABR) (Silver et al.. 2016; Alvarenga et al.. 2015; Pawlas et al.. 2015). In an unadjusted descriptive
analysis of children (4-13 years) in Poland, BLLs were positively correlated with brainstem auditory
evoked potentials (BAEP) and pure-tone audiometry and negatively correlated with acoustic otoemission
(Pawlas et al.. 2015). In multivariable analyses stratified by polymorphisms in the ALAD and vitamin D
receptor (VDR) genes, the associations for BAEP per (ig/dL increase in blood Pb were generally null
(Pawlas et al.. 2015). Silver et al. (2016) measured ABR in newborns (average 2 days old) in China.
Compared with a low (<2 (ig/dL) BLL measured during late pregnancy, infants exposed to medium (2-
3.8 (ig/dL) and high (>3.8 (ig/dL) Pb levels were more likely to have a higher ABR central-to-peripheral
(C-P ratio) (Silver et al.. 2016). When using Pb levels measured in cord blood and during mid-pregnancy,
however, the association for ABR C-P ratio moved toward the null (Silver et al.. 2016). Although
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quantitative results were not provided for a study of children (18 months-14 years) in Brazil, Alvarenga
et al. (2015) observed no association between cumulative BLLs and BAEP.
Summary
A prospective study in the 2013 Pb ISA (Dietrich et al., 1992) found an association of Pb
exposure with decreased auditory processing. In addition, cross-sectional studies found increased hearing
thresholds in children aged 4-19 years that participated in NHANES and HHANES in association with
higher concurrent BLLs. Recent cross-sectional and case-control studies of young children and
adolescents generally support a positive association between Pb exposure and hearing loss (Xu et al„
2020; Choi and Park, 2017; Shargorodskv et al., 2011), whereas the results were inconsistent for ABR
(Silver et al„ 2016; Alvarenga et al„ 2015; Pawlas et al„ 2015).
3.5.6.1.2 Toxicological Studies of Auditory Function
Toxicological evidence for effects on auditory function in the 2013 Pb ISA was limited to one
study (U.S. EPA, 2013). This study evaluated auditory thresholds using a behavioral task in 13-year-old
monkeys (Macaca mulatto) who had previously been exposed to Pb either gestationally or postnatally
(Laughlin et al., 2009). Potentially due to limitations noted within the study, small but nonsignificant
increases in the auditory threshold were reported in Pb-exposed animals compared with controls. Stronger
associations between Pb exposure, auditory threshold shifts, and latency in BAEP were reviewed in the
2006 Pb AQCD (U.S. EPA, 2006). Importantly, the associations demonstrated in the 2013 Pb ISA and
2006 Pb AQCD occurred at higher BLLs (>30 (ig/dL) that would not be considered PECOS-relevant for
this ISA.
Changes in auditory thresholds using BAEP have been further assessed in three recent rodent
studies (Table 3-16T). Jamesdaniel et al. (2018) exposed male C57B1/6 mice from PND 33 to PND 61 to
Pb and subsequently detected 8-12-dB upward shifts in hearing thresholds (indicative of hearing loss)
between 4 and 32 kHz. In contrast, another recent study using similarly aged male CBA/CaJ mice and a
longer exposure paradigm (11 weeks) found no significant effect of Pb on hearing thresholds at 8, 16, and
32 kHz (Carlson et al., 2018). The final study, which exposed male and female Sprague Dawley rats
postnatally to Pb did not detect significant differences in hearing thresholds between 4 and 28 kHz at
PND 60 (Zhu et al., 2016). Animals in the two studies that did not detect an effect had lower BLLs than
those in the study that did (3-8 (ig/dL versus 29 (ig/dL). However, due to the small number of studies, the
existence of an exposure threshold for this effect remains uncertain.
Recent studies have also investigated the effect of Pb exposures on auditory processing, which
was not discussed in previous ISAs. Zhu et al. (2016) exposed rat pups to Pb through their dams" drinking
water until weaning, when they began drinking Pb-free water. BLLs of the pups were roughly 8 (ig/dL
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during exposure and had returned to baseline levels by PND 40. At PND 60, the Pb-exposed rats were
found to have a decreased ability to discriminate between target and nontarget sound bursts. Additionally,
these rats were found to have a reduced spike rate-following ability and decreased cortical response
synchronization, indicative of a deficit in auditory cortical temporal processing. The same research group
published a follow-up study using a similar exposure paradigm to investigate another aspect of auditory
processing (i.e., sound localization) (Liu et al.. 2019). In a sound-azimuth discrimination task, Pb-exposed
animals took significantly longer to reach target accuracy and had significantly greater deviations (i.e.,
difference between the location of the desired response versus the location of incorrect response)
compared with control animals. These behavioral impairments were accompanied by a degraded sound-
azimuth selectivity in the primary auditory cortex neurons.
Summary
Earlier experimental animal studies have found decreased auditory function in adult monkeys and
rodents after lifetime exposure to Pb in animals with peak BLLs greater than 30 (ig/dL, but the persistence
of these effects at lower BLLs and in juvenile animals was uncertain. Three recent studies evaluated
auditory thresholds using BAEP in rodents exposed to Pb starting in the postnatal or juvenile period.
Jamesdaniel et al. (2018) found 8-12-dB upward shifts in hearing thresholds between 4 and 32 kHz in
young adult mice (peak BLLs of 29 (ig/dL). Studies evaluating lower mean BLLs from 3 to 8 (ig/dL did
not report differences in BAEP thresholds. However, mice with mean peak BLLs of 8 (ig/dL had
significant deficits in auditory processing, including decreased sound discrimination and sound
localization ability paired with dysfunction in the auditory cortical neurons (Liu et al.. 2019; Zhu et al..
2016).
3.5.6.2 Visual Function
The evidence reviewed in the 2013 Pb ISA was inadequate to determine whether a causal
relationship exists between Pb exposure and visual function in children (U.S. EPA. 2013). A study in
children and a few studies in animals showed Pb-associated increases in supernormal electroretinograms;
however, the biological plausibility of the observations was unclear. Overall, the available epidemiologic
and toxicological evidence was of insufficient quantity, quality, and consistency to support a causality
conclusion.
3.5.6.2.1 Epidemiologic Studies of Visual Function
Only a few epidemiologic studies examined the association between Pb exposure and decrements
in visual function in children (Silver et al., 2016; Fillion et al.. 2013). Since the measures of visual
function differed between studies, it is difficult to draw any conclusions about Pb exposure and visual
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function in children. Silver et al. (2016) measured grating visual acuity (VA) in 6-week-old infants in
China. Compared with low (<2 (ig/dL) BLLs measured during late pregnancy, infants exposed to medium
(2-3.8 (ig/dL) and high (>3.8 (ig/dL) Pb levels were more likely to have lower grating VA (Silver et al..
2016). When using Pb levels measured in cord blood and during mid-pregnancy, however, the association
for grating VA was attenuated and moved closer toward the null (Silver et al., 2016). In Brazil, Fillion et
al. (2013) measured contrast sensitivity (cycles per degree [cpd]) and acquired color vision loss (color
confusion index, CCI) in study volunteers that included adolescents (age range: 15-66 years). Based on
the entire study population, blood Pb exposure was negatively associated with the intermediate spatial
frequency of contrast sensitivity (12 cycles/degree); however, results varied by spatial frequency (Fillion
et al., 2013). For CCI, there was a small positive association with blood Pb (Fillion et al., 2013).
Summary
Overall, the available epidemiologic and toxicological evidence assessed in the 2013 Pb ISA was
of insufficient quantity, quality, and consistency to support a causality conclusion. A limited number of
recent epidemiologic studies are available for consideration; however, measures of visual function
differed between studies limiting observations regarding the consistency of the evidence overall.
3.5.6.2.2 Toxicological Studies of Visual Function
The evidence base pertaining to effects on visual function in the 2013 Pb ISA was largely
supported by seminal literature reviewed previously in the 1986 and 2006 Pb AQCDs showing reduced
VA, retinal alterations, and changes in CNS visual processing areas and subcortical neurons involved in
vision (U.S. EPA, 2013, 2006, 1986). Electroretinography (ERG), which measures the bioelectrical
response of the retina to a light stimulus, is used to detect abnormalities in retinal functioning. Fox et al.
(2008) found that Pb exposure in female Long-Evans rats (gestation through PND 10, measured at PND
90) induced supernormal ERGs (i.e., increases in the response amplitude) at low and moderate exposure
levels (BLLs of 12 and 24 (ig/dL) and subnormal ERGs (i.e., decreases in the response amplitude) in the
high exposure group (BLL of 46 (ig/dL). Earlier studies have also found Pb-related aberrations in ERGs,
but the direction of this effect is inconsistent (i.e., both subnormal and supernormal responses have been
detected) (Fox et al.. 1997; Lilienthal et al.. 1988). As discussed in Giddabasappa et al. (2011). the effect
direction may be related to both the lifestage during exposure (gestational versus postnatal) and the Pb
dose. This study also demonstrated that low to moderate gestational Pb exposure (BLLs: 10 and 27
(ig/dL) increased and prolonged retinal progenitor cell proliferation, resulting in selectively increased rod
photoreceptor and bipolar cell neurogenesis in C57BL/6 mice at PND 60 (Giddabasappa et al.. 2011).
Adult monkeys {Macaca fascicidaris) with lifetime Pb exposure, producing BLLs from 50-115 (ig/dL,
had temporal vision dysfunction but no change in spatial function (Rice. 1998). In contrast to these
effects, Laughlin et al. (2008) found that Pb exposure in Rhesus monkeys (exposed from PND 8-26
weeks; BLLs of 35-40 (ig/dL) did not significantly affect the development of photopic spatial acuity
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assessed using a modified Teller preferential looking paradigm. Recent PECOS-relevant studies have not
further examined the effects of Pb on visual function.
3.5.6.3 Relevant Issues for Interpreting the Evidence Base
3.5.6.3.1 Potentially At-Risk Populations
Genes
Pawlas et al. (2015) conducted multivariable analyses stratified by polymorphisms in the ALAD
and VDR genes and found that the associations for BAEP and pure-tone audiometry per |ig/L increase in
blood Pb were generally null (Pawlas et al.. 2015).
3.5.6.4 Summary and Causality Determination: Sensory Organ Function
The 2013 Pb ISA presented two causality determinations related to sensory function in children:
auditory function and visual function (U.S. EPA, 2013). At the time, the evidence was sufficient to
conclude that a causal relationship was likely to exist between Pb exposure and auditory function
decrements in children. For visual function, the evidence was inadequate to determine if a causal
relationship exists. In 2015, the Preamble to the ISA introduced minor changes to the language used in the
causality framework descriptors (U.S. EPA, 2015). This change has affected the causality determination
for this section. Importantly, the new determination is not intended to be interpreted as a weakening of the
evidence base, as recent evidence has remained consistent with previously reviewed studies.
Auditory processing decrements were previously demonstrated in a prospective study by Dietrich
et al. (1992). In 5-year-old children, elevated BLLs during infancy (mean BLLs of 4.8 (ig/dL at 10 days
old) were associated with poorer performance on a test for auditory processing disorders after adjusting
for confounding factors including SES, HOME score, a variety of birth outcomes, maternal alcohol
consumption, maternal smoking, and overall child health. Recently, experimental animal studies
demonstrated that postnatal Pb exposure resulting in mean peak BLLs of 8 (ig/dL also caused significant
deficits in auditory processing, including decreased sound discrimination and sound localization ability
paired with dysfunction in the auditory cortical neurons (Liu et al„ 2019; Zhu et al„ 2016).
Multiple large cross-sectional NHANES and HHANES studies have shown that higher BLLs
(children aged 4-19; BLLs 8 (ig/dL) are associated with increased hearing thresholds (Schwartz and Otto,
1991, 1987). These associations remained after adjustment for age, sex, race, family income, parental
education, and nutritional factors. Recent cross-sectional and case-control studies continued to
demonstrate associations with BLLs and hearing loss in young children (aged 3-7, BLLs ~3 to 6 (ig/dL)
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and adolescents (aged 12-19, BLLs ~1 to 8 (.ig/dL). particularly at higher frequencies (Xu et al.. 2020; Liu
et al., 2018c; Choi and Park, 2017; Shargorodskv et al., 2011). Furthermore, hearing threshold increases
were previously demonstrated in adult nonhuman primates after developmental or lifetime Pb exposure,
although BLLs in these studies were greater than 30 (ig/dL (Laugh 1 in et al., 2009; Rice, 1997). Recent
experimental animal studies have not further evaluated hearing thresholds in nonhuman primates and
instead have focused on BAEPs in rodents. Jamesdaniel et al. (2018) found 8-12-dB upward shifts in
auditory thresholds between 4 and 32 kHz in young adult mice exposed during adolescence (peak BLLs
29 (ig/dL). Similar studies did not detect differences in BAEPs in rodents with lower peak BLLs (3 to 8
(ig/dL). Likewise, a few recent epidemiologic studies also evaluated BAEP with inconsistent results.
Although Pb-induced alterations in subcortical visual neurons, visual processing areas, and retinal
development have been demonstrated, supporting the biological plausibility of Pb-associated effects on
vision (U.S. EPA, 2013), evidence relating to visual function in epidemiological and toxicological studies
remains limited and inconsistent. Silver et al. (2016) found that decreased visual acuity in infants was
associated with maternal BLLs higher than 2 (ig/dL in late pregnancy, but this association was weaker
with BLLs in both mid-pregnancy and cord blood. Studies in nonhuman primates failed to detect changes
in visual acuity at BLLs above 35 (ig/dL, although one reported decrements in temporal acuity as a result
of Pb exposure (Laugh 1 in et al., 2008; Rice, 1998). Another recent study found associations with blood Pb
and decrements in contrast sensitivity and color vision, an endpoint that has not been previously studied,
in a study population that included adolescents (15-66 years old) (Fillion et al., 2013). Studies in both
humans and animals have found significant but inconsistent changes in ERGs (Fox et al., 2008;
Rothcnbcrg et al., 2002; Fox et al., 1997), though it is unclear if these findings translate to functional
visual changes.
In conclusion, the evidence is suggestive of, but not sufficient to infer, a causal relationship
between Pb exposure and sensory function in children. This determination is based primarily on the
strongest line of evidence within the sensory function grouping {i.e., auditory function). No recent
epidemiologic studies have further investigated the auditory processing decrements shown in Dietrich et
al. (1992), but recent experimental animal studies have demonstrated Pb-induced effects on auditory
processing. Cross-sectional and case-control studies focusing on the impact of Pb exposure on hearing
loss generally support an association but are not entirely consistent. Experimental animal studies
evaluating hearing loss at human relevant BLLs in young animals are not available. Limited
epidemiologic studies have evaluated Pb exposure and visual function in children with inconsistent
findings, but evidence for biological plausibility has been demonstrated.
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Table 3-7 Evidence that is suggestive of, but not sufficient to infer, a causal relationship between Pb
exposure and sensory organ function in children
Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with
Effects0
Auditory Function
Consistent findings from a few Prospective study found associations of Dietrich et al. (1992)
epidemiologic studies with prenatal (maternal), neonatal, yearly age
relevant BLLs 1 to 5 yr, lifetime avg BLLs with poorer
auditory processing in children at age 5 yr
in Cincinnati.
Blood Pb means: neonatal (10 d) 4.8
|jg/dL, yearly age 1 to 5 yr 10.6-17.2
|jg/dL, lifetime (to age 5 yr) avg NR
Cross-sectional and case-control studies
for increased hearing thresholds in
children ages 3-19 yr, including analyses
of NHANES, HHANES and KNHANES in
association with higher concurrent BLLs.
Section 4.3.6.1, (U.S. EPA, 2013)
Xu et al. (2020)
Liu et al. (2018c)
Blood Pb median:
HHANES: 8 pg/dL; NHANES: NR
Means 3.63-5.69 pg/dL (3-7 yr)
Sharqorodsky et al. (2011)
NHANES (2005-2008): med ~1 pg/dL
(12-19 yr)
Choi and Park (2017)
KNHANES: GM: 1.26 pg/dL (15.6 yr)
Epidemiologic evidence helps
to rule out chance, bias and
confounding with reasonable
confidence
Prospective study adjusted for SES,
HOME score, birth outcomes, obstetrical
complications, maternal smoking. Several
other factors considered.
Dietrich et al. (1992)
Cross-sectional and case-control studies
considered potential confounding by age,
sex, race, income, parental education,
nutritional factors.
Xu et al. (2020)
Liu et al. (2018c)
Sharqorodsky et al. (2011)
Choi and Park (2017)
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Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with
Effects0
Uncertainty due to lack of
animal evidence in juveniles
and limited evidence at
relevant exposure levels
Hearing loss in adult monkeys and
decreased BAEP in young adult rodents
at higher exposure levels.
Rice (1997)
Lauahlin et al. (2009)
Jamesdaniel et al. (2018)
Peak BLLs >29 |jg/dL
Decrements in sound discrimination and
localization in young adult rodents.
Zhu et al. (2016)
Liu et al. (2019)
Peak BLLs 8.2 pg/dL
Visual Function
Limited evidence from
epidemiologic studies
Associations with some tests of grating
VA and contrast sensitivity observed.
Silver et al. (2016)
Fillion et al. (2013)
Uncertainty due to limited
animal evidence in juveniles
and at relevant exposures
Higher than relevant postnatal Pb
exposure did not cause changes in VA in
infant nonhuman primates in infants but
did decrease temporal acuity in adults.
Lauahlin et al. (2008)
Rice (1998)
BLLs >35 pg/dL
Biological plausibility
demonstrated
Pb-induced alterations in ERGs,
subcortical visual neurons, visual
processing areas, and retinal
development demonstrated.
(U.S. EPA, 2013)
avg = average; BAEP = brainstem auditory evoked potentials; BLL = blood lead level; d = day; ERG = electroretinography; GM = geometric mean; HHANES = Hispanic Health and
Nutrition Examination Survey; HOME = Health Outcomes and Measures of the Environment; KNHANES = Korea National Health and Nutrition Examination Survey; NHANES =
National Health and Nutrition Examination Survey; NR = not reported; Pb = lead; SES = socioeconomic status; VA = visual acuity; yr = year(s).
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 20151. Note that
the change from "likely to be causal" for auditory effects in children in the 2013 Lead ISA, to "suggestive of, but not sufficient to infer, a causal relationship" for sensory organ function
in children reflects minor changes to the causal framework, rather than a weakening of the evidence base pertaining to auditory effects in children.
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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3.5.7
Social Cognition and Behavior in Children
In addition to neurodevelopmental disorders covered in previous sections—including ADHD,
intellectual and developmental disabilities, and motor disorders—there is an emerging body of research
on autism spectrum disorder (ASD) and other conditions related to social cognition and behavior. The
2013 Pb ISA (U.S. EPA, 2013) did not evaluate any epidemiologic studies examining associations
between Pb exposure and autism. ASD is generally characterized by restricted interests and behaviors,
including stereotyped patterns of behavior and sensory sensitivities. To meet the DSM criteria for ASD, a
child must have persistent deficits in social communication and demonstrate repetitive behaviors (APA,
2013). Social cognition, which is often impaired among individuals with ASD, involves the ability to
interpret and respond to social cues, communication, and interaction. These traits (or behaviors) can be
measured on a continuum in the general population with scores exhibiting a fairly normal distribution,
with scores at the extreme impaired end indicating a higher risk for ASD (Constantino, 2011). Deficits in
social cognition have been associated with lifelong educational, vocational, adaptive functioning, and
mental health challenges among individuals with and without a clinically diagnosed disorder. Autism
diagnosis (e.g., via the ICD code), the CBCL, Social Responsiveness Scale (SRS), BASC-2, BSID-II and
III, CDIIT, ASQ:I, GDS, Social Maturity Scale (SMS), MDAT, and ECDI have been used in studies
examining the association of Pb exposure with social cognition and behavior.
3.5.7.1 Epidemiologic Studies of Social Cognition and Behavior
There have been a number of recent studies of ASD and deficits in social cognition and related
behaviors. Many of these recent studies did not control for potential confounders and/or did not include
robust statistical methods to estimate C-R relationships between Pb exposure and outcome, and are not
considered further in this section (Filon et al.. 2020; Qin et al.. 2018; Skalnv et al.. 2017; Macedoni-
Luksic et al.. 2015; Alabdali et al.. 2014; Yassa. 2014; De Palma et al.. 2012; Blaurock-Busch et al..
2011; Tian et al.. 2011). Instead, the ensuing discussion focuses on a number of autism and social
cognition studies that include more comprehensive control for potential confounders. The relevant studies
provide some evidence of a positive association between Pb exposure and ASD, along with generally
consistent supporting evidence of an association with decrements in social cognition. Measures of central
tendency for BLLs used in each study, along with other study-specific details, including study population
characteristics and select effect estimates, are highlighted in Table 3-13E of Section 3.7. An overview of
the recent evidence is provided below.
Two recent studies used robust modeling approaches to assess the C-R relationship between
exposure to Pb and ASD (Arora et al.. 2017; Kim et al.. 2016). While each study examined different
biomarkers of exposure and exposure windows, both indicated associations between Pb exposure and
ASD. Arora et al. (2017) conducted a difference-in-differences analysis of a small case-control study of
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8- to 12-year-old twins with discordant or concordant ASD status. Pb was measured in shed deciduous
teeth using a method that provided temporal estimates of tooth Pb levels ranging from 20 weeks before
birth to 30 weeks after birth. To estimate the relationship between tooth Pb and ASD across this exposure
window, the authors used distributed lag models to estimate the smoothed mean differences in tooth Pb
levels in discordant pairs minus the mean differences in concordant twins at each time point. In this case,
concordant twins served as the control group to account for natural variations in Pb exposure within a
dyad. One analysis used concordant twins without ASD and the other used concordant twins with ASD as
the control groups. In both cases, the difference in tooth Pb levels between discordant twins was greater
than the difference in concordant twins across the entire exposure window, though there appeared to be
bimodal peaks in tooth Pb differences from about 10 to 15 weeks before birth and 10 to 20 weeks after
birth (see Figure 3-14).
Time since birth (weeks) Time since birth (weeks)
ASD = autism spectrum disorder.
Black line represents the difference in mean differences in tooth Pb levels between discordant ASD twins and: A) control twins; or B)
concordant ASD twins. Gray bands are unadjusted 95% CIs, while blue bands are adjusted for intra-twin correlations. Values above
zero represent increased levels in ASD cases compared with the non-ASD sibling after taking into account average difference in
control twins.
Source: Arora et al. (20171.
Figure 3-14 Differences in mean difference tooth Pb levels for autism
spectrum disorder in discordant twin pairs versus (A) non-autism
spectrum disorder twin pairs or (B) autism spectrum disorder
concordant twin pairs.
In a large cohort study of children in South Korea, Kim et al. (2016) analyzed blood Pb in relation
to autistic behaviors measured by parental response to the Autism Spectrum Screening Questionnaire
(ASSQ) and the SRS at ages 11-12 years old. BLLs at study enrollment (7-8 years old) were associated
with higher scores on the ASSQ (number of autistic behaviors) and SRS (severity across domains of
social awareness, cognition, communication, motivation, and mannerisms). There were null associations
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with blood Pb measured at 9-10 years and attenuated, but still positive, associations with concurrent
BLLs (11-12 years old). Nonparametric generalized additive models indicated an approximately linear
relationship between BLLs at enrollment and scores on the SRS. In addition to continuous models, the
authors also dichotomized ASSQ scores and reported 45% higher odds (95% CI: 10%, 93%) of a positive
screen for autism (ASSQ score >17) per 1 (ig/dL higher BLL at enrollment. Notably, symptoms of ASD
manifest as early as infancy and although BLLs in this study were measured prior to assessment of
autistic behaviors, the relevant exposure window likely preceded exposure measurement.
In contrast to the results from Kim et al. (2016) and Arora et al. (2017). recent case-control
studies did not observe an association between adjusted mean childhood BLLs and ASD cases (Rahbar et
al.. 2021; Rahbar et al.. 2015). In addition to matching cases and controls on age and sex, the authors
estimated mean differences using a linear model controlling for a variety of demographic and SES factors,
including maternal age. Similarly, a recent case-control study reported a null association between tertiles
of maternal BLLs and ASD in children, though there was some evidence of a nonlinear association in a
cubic spline model (Skogheim et al.. 2021). The study populations for these analyses included children
and mothers with lower (<2 (ig/dL; (Skogheim et al.. 2021; Rahbar et al.. 2015)) and higher (>7 (ig/dL;
(Rahbar et al.. 2021)) mean or median BLLs.
One additional large retrospective study in Northeast China (Dong et al.. 2022) compared current
BLLs in children with moderate/severe versus mild autism, as determined by CARS scores. Mean BLLs
for the mild and moderate/severe groups were 2.58 (SD: 1.08) (ig/dL and 3.25 (SD: 1.89) (ig/dL,
respectively. After adjusting for age, residence, parental caregiving, parental education, and
gastrointestinal conditions, autism severity was positively associated with BLL (|3 = 0.03 [95% CI: 0.01,
0.05]).
Additional supporting evidence was provided by several cohort studies that investigated
associations between Pb exposure and social cognition in children without autism. Most of these studies
reported inverse relationships between prenatal Pb exposure and social cognition measures. Several
studies additionally investigated effect modification by various other factors.
Rvgiel et al. (2021) assessed the relationship between maternal blood Pb and infant behavioral
development at 12 to 24 months of age in a small analysis of three birth cohorts from the ELEMENT
study in Mexico City. The authors used the behavioral rating scale (BRS) of the BSID-II to examine
attention, social engagement, orientation, motivation, and emotional response, giving rise to two social
cognition outcomes: orientation/engagement (ORIEN) and emotional regulation (EMOCI). The authors
reported that children had lower 24-month EMOCI and ORIEN percentile ranks with higher maternal
BLLs. Associations were observed in relation to maternal BLLs measured during each trimester, but
greatest for the second trimester, with 1.13% (95% CI: -2.63%, 0.37%) and 0.98% (95% CI: -2.83%,
0.88%) lower 24-month EMOCI and ORIEN percentiles, respectively, for 1 (ig/dL higher second
trimester BLLs. In an examination of the mediation of trimester-specific Pb exposure by DNA
methylation at several previously identified CpG sites, the authors observed both enhancing and
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suppressive effects of DNA methylation on the association between blood Pb and neurocognitive
outcomes, depending on the gene locus, with methylation at the majority of loci playing a suppressive
role. Shekhawat et al. (2021) similarly reported null but slightly inverse associations between cord blood
Pb and social-emotional scores from the BSID-III in a study of mother-child pairs in western Rajasthan,
India. Neurocognitive assessments were conducted at the average age of 6.5 months. Additionally, in an
analysis of the Navajo Birth Cohort Study, Nozadi et al. (2021) reported imprecise inverse associations
between maternal BLLs and communication (|3 = -0.15 [95% CI: -0.58, 0.28]) and personal-social (|3 =
-0.11 [95% CI: -0.72, 0.50]) domain scores on the ASQ:I at 10 months.
Some cohort studies examined interactions between Pb levels and the levels of other trace
elements (Nvanza et al.. 2021; Dohertv et al.. 2020; Lin et al.. 2013). Lin et al. (2013) measured maternal
blood Pb and assessed child development (including social and self-care skills) in the TBPS with the
CDIIT, as described in Section 3.5.1.2. The authors observed that children with high Pb exposure (>75th
percentile: 1.65 (ig/dL) had lower social DQs (|3 = -5.89 [95% CI: -10.81, -0.97]) compared with those
with low prenatal Pb exposure. In addition, the authors reported lower social or self-help DQs among
those with higher Pb and Mn concentrations in an interaction analysis. Nvanza et al. (2021) measured Pb,
Hg, Cd, and As concentrations using dried blood spots from pregnant mothers at 16-27 weeks of
gestation in Northern Tanzania. Adjusting for maternal age, maternal education, maternal and parental
occupation, number of under-five siblings at home, family socioeconomic wealth quintile, infant sex,
infant age, birth weight, and height and weight at neurocognitive testing, the authors did not observe an
association between high Pb exposure (>3.5 (ig/dL) and social impairment on the MDAT, which is
described in Section 3.5.1.2. However, an interaction analysis with maternal blood Hg levels showed that
children highly exposed to both Hg (>0.08 (ig/dL) and Pb were more likely to have global
neurodevelopmental impairment (PR= 1.40 [95% CI: 0.90, 2.10]). Dohertv et al. (2020) measured
concentrations of Pb and other metals (As, Cu, Mn, Se, and Zn) in maternal prenatal and postnatal
toenails and infant toenails at 6 weeks of life from mother-infant pairs in the New Hampshire Birth
Cohort. The three exposure assessments estimated exposures that occurred during periconception and
early pregnancy, mid-pregnancy, and late pregnancy and early neonatal life, respectively. The authors
observed mostly negative but imprecise associations between prenatal and child toenail Pb levels and total
SRS-2 scores. They also observed mostly positive but imprecise associations between postnatal maternal
and child toenail Pb levels and the adaptive skills composite on the BASC-2 (see Section 3.7, Table
3- 13E). Pb concentrations did not appear to interact with other metals on the total SRS-2 score or the
BASC-2 adaptive skills composite, and sex-stratified analyses revealed inconsistent associations among
girls.
An additional study assessed effect modification by maternal psychosocial measures. Zhou et al.
(2017) investigated the interactions of maternal BLL in whole blood and maternal prenatal stress levels
with child development (including adaptive behavior and social domains) using the GDS. Among those
with high maternal stress levels (GSI: P75-P100), adaptive behavior DQs were 17.93 points lower (95%
CI: -35.83, -0.03) per loglO-transformed (ig/dL higher maternal BLL. Social behavior DQs were also
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inversely associated with maternal BLL in children of mothers with high stress levels (|3 = -41.00 [95%
CI: -63.11, -18.89] per log-10 transformed unit of BLL).
One cross-sectional study (Ruebner et al.. 2019) evaluated the association between concurrent
BLL and neurocognitive outcomes including adaptive skills among children with CKD, using parent
ratings on the BASC-2. This study is discussed in more detail in Section 3.5.2.1.1. Higher BLL was
associated with worse adaptive skills composite scores (|3 = —3.1) in univariable analyses; however, this
association did not remain after adjusting for key sociodemographic and clinical confounders.
One additional prospective study (Vigeh et al.. 2014) measured domains of social and self-help
skills but presented only associations with composite neurodevelopment test scores (described in Section
3.5.1.2), which impedes parsing of specific social cognition effects of Pb exposure. In addition, Kim et al.
(2018b) evaluated concentrations of Pb in maternal serum, cord blood, urine, and breast milk in
association with neurodevelopmental and behavioral outcomes, including social quotient (SQ) measures
from the SMS among 13-24-month-old children. However, they reported only statistically significant
results in the paper, precluding quantitative results for blood Pb and SQ.
Recent epidemiologic studies utilized a wide range of outcome measures, including diagnostic
tests of autism (e.g., ICD code, DSM classification, ASSQ, and the Autism Diagnostic Observation
Schedule [ADOS]), behavior rating systems (e.g., CBCL, SRS-2, BASC-2, and SMS), and
neurodevelopmental assessments with social behavioral subtests (e.g., BSID-II and III, CDIIT, ASQ:I,
GDS, MDAT, and ECDI). Psychometric tests of social cognition often add valuable dimensional
information regarding the severity and type of social deficit among children with autistic traits, and the
wide variety of tests used in the evaluated studies provided insight into diverse aspects of problems with
social cognition, including communication, adaptive and self-help skills, social engagement, and
emotional behavior. One limitation, however, is that this variety complicates a straightforward
interpretation of results due to the lack of consistency of measures. Many behavioral tests provide
outcomes that overlap with domains discussed in other sections such as externalizing behavior (Section
3.5.3) and internalizing behavior (Section 3.5.4), which can limit parsing of effects. Vigeh et al. (2014)
examined social and adaptive skills but reported quantitative results using only the global
neurodevelopmental composite score from the ECDI. Other studies (Nvanza et al.. 2021; Rvgicl et al..
2021) used rating subscale measures (i.e., EMOCI and ORIEN from the BRS; social development score
from the MDAT) that are not widely used in the literature, making it difficult to compare results across
studies.
3.5.7.1.1 Summary
Two recent high-quality studies of Pb exposure and ASD reported positive associations between
increased Pb exposure and higher risk of ASD diagnosis or symptomatology (Aroraet al.. 2017; Kim et
al.. 2016). One retrospective study also observed a positive association between greater autism severity
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and current BLL (Dong et al.. 2022). However, some case-control studies did not find evidence of a
positive association (Rahbar et al.. 2021; Skogheim et al.. 2021; Rahbar et al.. 2015). Although most
autism studies except one (Kim et al.. 2016) were case-control studies, two of the case-control studies
accounted for temporality of exposure and outcome by analyzing prenatal maternal blood (Skogheim et
al.. 2021) or using tooth Pb measurement methods that allow ascertainment of perinatal Pb exposure
levels (Arora et al.. 2017). Several cohort studies observed null or slight impairments of social
dimensions scores. Recent studies of social cognition in children without ASD used a wide variety of
psychosocial and neurodevelopmental instruments, such as BASC-2, BSID-II and III, CDIIT, ASQ:I,
GDS, SRS, SMS, MDAT, and ECDI, to obtain scores of social, emotional, and adaptive abilities. These
studies were mostly prospective in design and accounted for some key potential confounders, including
maternal age, parental education, SES, and caregiving.
3.5.7.2 Toxicological Studies of Social Cognition and Behavior
The previous ISA incorporated evidence of the effects of Pb exposure on social cognition and
behavior. Donald et al. (1986) reported sex-specific effects of Pb exposure on social investigatory
behavior in mice, wherein males and females exposed to Pb displayed enhanced social interaction but at
different times after exposure. In a subsequent publication, Donald et al. (1987) reported that Pb exposure
increased non-social behavior in males while females displayed decreased non-social behavior. The
previous evidence suggests that Pb may influence social behavior in rodents in a sex-specific manner, but
the direction of the effect was not clear.
There is limited recent toxicological evidence available on the effects of Pb exposure on social
cognition and behavior. A single study by Tartaglione et al. (2020) examined homing test and ultrasonic
vocalizations (USV). USV are calls emitted by pups when separated from their mother and siblings and
are markers of early emotional and communication development. Pups prenatally and lactationally
exposed to Pb exhibited reduced numbers of calls at PND 4 and 12, with no significant differences at
PND 7 and 10 from control animals. The same study also performed a homing test, which assesses
discriminative performance and maternal preference behavior by separating the pup from the dam and
recording the time taken to return to the nest from a maze. The time spent is a measure of both olfactory
discrimination and social preference. The authors reported no difference in homing test performance
between control and Pb-exposed pups at PND 12 (Tartaglione et al.. 2020). In summary, there is limited
evidence from the toxicological literature examining potential relationships between developmental Pb
exposure and social behavior, which represents an area of uncertainty.
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3.5.7.3 Relevant Issues for Interpreting the Evidence Base
3.5.7.3.1 Concentration-Response Function
Evaluation of the shape of the C-R function in recent studies of social cognition is limited,
making it challenging to draw conclusions. Across studies, associations between Pb exposure and social
cognition and behavior were observed at median or geometric mean maternal and cord BLLs ranging
from 3.3 to 5.5 (ig/dL, and BLLs measured in children ranging from 1.6 to 3.9 (ig/dL (Table 3-8). Kim et
al. (2016) used penalized regression splines to examine the C-R relationship between BLLs at 7-8 years
old and SRS scores at 11-12 years old. The C-R relationship was approximately linear across the range of
the BLL distribution, though there is more confidence in the shape of the C-R relationship (i.e., more
narrow confidence limits) closer to the mean, where there is a higher density of observations. Spline
models for most of the SRS subscales are also approximately linear, except for social cognition, which
has a sublinear relationship with BLLs (i.e., a smaller slope below the mean).
3.5.7.3.2 Potentially At-Risk Populations
Maternal Stress
There is limited evidence that maternal stress modifies the association between Pb exposure and
social cognition. Stratifying by maternal stress, Zhou et al. (2017) found that social behavior (|3 = -41.00,
95% CI: -63.11, -18.89 per log-10 transformed unit of BLL) and adaptive behavior (|3 = -17.93, 95%
CI: -35.83, -0.03 per log-10 transformed unit of BLL) in toddlers were inversely associated with BLLs
among children of mothers with high prenatal stress. In contrast, adaptive behavior appeared to have a
positive but imprecise relationship (|3 = 7.57, 95% CI: -0.12, 15.27 per log-10 transformed unit of BLL)
with BLLs among children of mothers with low prenatal stress, while the association with social behavior
was null in the same population.
Co-exposure to Other Metals or Chemicals
A limited number of studies examined co-exposures to other metals as potential modifiers of the
relationship between Pb and social cognition. Lin et al. (2013) observed slight impairments to social and
self-help DQs among those with high concentrations of both Pb (>1.65 (ig/dL) and Mn (>5.93 (.ig/dL).
Nvanza et al. (2021) conducted interaction analyses of Pb and various neurodevelopmental outcomes with
Hg, Cd, and As, but did not report results for social skills.
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Gene-Environment Interactions
A single study evaluated the role of DNA methylation as a mediator of the relationship between
Pb exposure and social cognition and behavior. Rygiel et al. (2021) found both enhancing and
suppressing effects of DNA methylation at several CpG sites in mediation analyses. Methylation of
cg23280166 within CCSER1, a gene which has been associated with ADHD, suppressed the association
between second trimester Pb levels and ORIEN and EMOCI scores at 24 months old, while methylation
at cgl8515027 (GCNT1), positively mediated the association between first and second trimester BLLs
and 24-month EMOCI scores. Likewise, DNA methylation of cg23280166 (VPS11) also positively
mediated the relationship between third trimester BLLs and 24-month EMOCI scores.
Pre-existing Conditions
Although no recent studies evaluated pre-existing conditions as potential effect modifiers, one
study evaluated the relationship between Pb exposure and adaptive behavior among children with CKD.
After adjusting for sociodemographic and CKD-related variables, they did not report quantitative results
because they did not observe a statistically significant association (Rucbncr et al.. 2019).
Sex
There is limited evidence on sex as a modifier of the association between Pb exposure and social
cognition and behavior. Dohertv et al. (2020) observed inconsistent associations between Pb and SRS-2
total and BASC-2 adaptive skills composite scores in sex-stratified analyses. Female infant toenail Pb
concentration was positively associated with adaptive skills (|3 = 0.26 [95% CI: 0.07, 0.45] per log-2
transformed unit of BLL) and maternal prenatal toenail Pb was negatively associated with adaptive skills
in female infants (|3 = -0.19 [95% CI: -0.34, -0.04] per log-2 transformed unit of BLL), but associations
in male infants were null. Sample size limited statistical precision in sex-stratified analyses, which may
help explain these inconsistencies.
3.5.7.3.3 Confounding
Several sociodemographic characteristics were considered as potential confounders in recent
epidemiologic studies. Child age at outcome measurement was included in all but three studies
(Shekhawat et al.. 2021; Kim et al.. 2018b: Vigeh et al.. 2014) and child sex was included in all studies
but two (Nozadi et al.. 2021: Vigeh et al.. 2014). Parental education, which was consistently associated
with BLLs and measures of social cognition and/or autism status, was adjusted for or considered in all
studies except Rygiel et al. (2021). However, Rygiel et al. (2021) was the only study to include maternal
IQ as a potential confounder. Many studies also included SES among their modeled covariates (Nvanza et
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al., 2021; Rygiel et al.. 2021; Ruebner et al., 2019; Zhou et al., 2017; Vigeh et al.. 2014). Quality of
parental caregiving (e.g., HOME score) was included in only one study (Lin et al., 2013).
Various pregnancy and birth factors are also relevant for consideration as potential confounders.
Maternal age is strongly associated with autism risk (Sandin et al„ 2012) and is correlated with Pb
exposure (Ettinger et al„ 2020); hence, lack of inclusion in models may introduce bias. Additionally,
autism, like many developmental disorders, is more prevalent as delivery diverges in both directions from
40 weeks of gestation. As such, gestational age, birth weight, and maternal age were consistently included
as a potential confounder in most analyses. Breastfeeding, parity, maternal smoking and alcohol intake,
and food consumption during pregnancy were also included in multiple studies.
Genetics may also play a large role in the association between Pb exposure and social cognition
abilities. Aroraet al. (2017) used a case-control design with twin pairs, which allowed for matching on
genetic factors to some extent. Rahbar et al. (2021) evaluated interaction effects of glutathione S-
transferase (GST) genes (GSTP1, GSTM1, and GSTT1), which have been linked to detoxification of
environmental pollutants and to autism status. Rygiel et al. (2021) examined mediation by DNA
methylation at various CpG sites linked to prenatal Pb levels.
Co-exposures and mixtures with other trace metals were considered in several studies. Nozadi et
al. (2021) found positive correlations of BLLs with Mn and Cd. The authors used an algorithm to identify
the control variables for each metal they analyzed, including all co-occurring metals and demographics;
however, none met the inclusion criteria of being significantly associated with both the exposure and
outcome, and the final model did not include any covariates. Additionally, Lin et al. (2013) and Nyanza et
al. (2021) reported that Pb was positively correlated with Mn and As, and Cd and Pb, respectively.
However, neither study adjusted for metals in their analyses. Kim et al. (2016) adjusted for Hg and was
the only study to adjust for a co-occurring metal in its final model.
3.5.7.3.4 Lifestages
No epidemiologic studies examining the relationship between Pb exposure and social cognition
and behavior in children were included in the 2013 Pb ISA (U.S. EPA, 2013). Recent studies
demonstrated that BLLs measured during various lifestages and time periods (i.e., prenatal, early
childhood, later childhood, and concurrent with outcome assessment) are associated with ASD and
decrements in social cognition. Due to differences in study designs and the variety of psychometric tests
used to assess aspects of social cognition, it is difficult to compare the magnitude of associations across
studies to characterize important lifestages and time periods of Pb exposure. There is some examination
of different exposure measurement windows within studies. In the case-control study of twins described
previously, Arora et al. (2017) used laser ablation-inductively coupled plasma-mass spectrometry (ICP-
MS) to estimate pre- and postnatal Pb exposure from shed deciduous teeth. Differences in tooth Pb levels
were consistently higher in discordant ASD twins across the exposure period (20 weeks prenatal to 30
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weeks postnatal) compared with concordant and control twins, with bimodal peaks around 10 to 15 weeks
before birth and 10 to 20 weeks after birth (see Figure 3-14). This is consistent with results from a birth
cohort study that reported negative associations between maternal BLLs and social cognition in infants
(Rvgiel et al.. 2021). The observed associations were strongest in magnitude with maternal BLLs
measured in the second trimester compared with BLLs in the first and third trimesters. Although the
limited number of studies that evaluate different exposure windows makes it difficult to draw firm
conclusions on critical lifestages, the nature of ASD as a developmental disorder suggests that prenatal
and early infant exposures may be of particular importance.
It should be noted that children with ASD have a high prevalence of pica, a compulsive eating
behavior of non-food items (Fields et al.. 2021). Thus, children with ASD may have elevated BLLs due to
their higher likelihood of ingesting soil or other materials contaminated with Pb, rather than Pb exposure
causing ASD. As there is potential for reverse causation, accurately ascertaining the time of exposure
measurement is crucial in order to determine whether a causal effect of Pb on ASD exists, and studies
with exposure metrics that precede pica behavior would mitigate this concern. Such metrics include bone
Pb, tooth Pb (Arora et al.. 2017). and cord or maternal blood Pb (Nozadi et al.. 2021; Nvanza et al.. 2021;
Rvgiel et al.. 2021; Shekhawat et al.. 2021; Skogheim et al.. 2021; Kim et al.. 2018b; Zhou et al.. 2017;
Vigeh et al.. 2014; Lin et al.. 2013).
3.5.7.4 Summary and Causality Determination: Social Cognition and Behavior
The 2013 Pb ISA (U.S. EPA. 2013) did not include a causality determination for social cognition
and behavior in children. There were no epidemiologic studies on social cognition and behavior in
children in the previous ISA, and only a few toxicological studies that examined social behavior in mice.
The number of studies examining autism and social cognition in relation to Pb exposure has increased
substantially since the 2013 Pb ISA (U.S. EPA, 2013), highlighted by recent epidemiologic studies that
provide some evidence that Pb exposure is associated with increased ASD incidence and symptomology,
as well as decrements in social, emotional, and adaptive abilities. Recent toxicological evidence, along
with studies reviewed in the 2013 Pb ISA, provide some evidence of Pb-induced changes in social
behavior in mice, but the direction of the observed changes was inconsistent.
A recent novel epidemiologic analysis of twins provides strong evidence of an association
between Pb exposure and ASD. Arora et al. (2017) examined tooth Pb levels with respect to ASD status
among discordant and concordant twin pairs and observed higher Pb levels in the affected twin among
discordant monozygotic and dizygotic pairs. In contrast, concordant twins demonstrated similar levels of
exposure. This study also provided some insight into potentially sensitive time windows of exposure in
which the association between tooth Pb levels and autistic status was highest between 10-15 weeks
before birth and 10-20 weeks after birth. Additional support was provided by a prospective cohort study,
which reported that the number and severity of autistic behaviors in young children was positively
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associated with low BLLs (geometric mean: 1.58-1.64 (ig/dL) at several points prior to outcome
assessment (Kim et al.. 2016). There is some uncertainty about the relevance of the exposure window in
this study given that the earliest Pb measurements occurred at 7-8 years old, which is close to the
outcome assessment age (11-12 years old) and later than autistic behaviors typically manifest.
Additionally, covariates examined in this study did not include maternal age, which is an important
potential confounder for developmental disorders like autism; therefore, lack of adjustment for this
variable weakens the conclusions that can be drawn from the analysis. Dong et al. (2022) provides
support for the positive association between autism severity and BLLs among children 2 to 13 years old at
low levels of current blood Pb (mild group mean: 2.58 (ig/dL; moderate/severe group mean: 3.25 (.ig/dL);
however, the study's retrospective design and the wide range of the ages of assessed children introduce
uncertainty regarding potential reverse causality.
Several prospective studies among children without autism provide some additional support for
associations between Pb exposure and measures of social impairment in children (Nozadi et al.. 2021;
Nvanza et al.. 2021: Rygiel et al.. 2021: Shekhawat et al.. 2021: Zhou et al.. 2017: Lin et al.. 2013).
Median or geometric mean maternal and cord BLLs in these studies ranged from 3.3 to 5.5 (ig/dL, and
BLLs measured in children ranged from 2.7 to 3.9 (ig/dL. These studies had moderate to good follow-up
participation rates, and follow-up durations ranged from 6.5 months to 3 years. Furthermore, they
demonstrated good confounder control, adjusting for maternal age and some measure of SES or parental
education. Notably, the use of non-specific composite test scores (Vigeh et al.. 2014) and lesser-used
subscales (Nvanza et al.. 2021: Rygiel et al.. 2021) limits the specificity and generalizability of some
studies. Additionally, results from recent studies were not entirely consistent, as some analyses did not
observe associations (Rahbar et al.. 2021: Skogheim et al.. 2021; Dohertv et al.. 2020; Ruebner et al..
2019; Rahbar et al.. 2015). These included mostly case-control studies, one prospective cohort study, and
one cross-sectional study. Although all three case-control studies adjusted for maternal age and various
relevant covariates among matched pairs, Ruebner et al. (2019) did not.
Two toxicological studies in the 2013 Pb ISA reported a potential sex-based effect modification
of the effect of Pb exposure on social behavior (Donald et al.. 1987. 1986). Female and male mice
exhibited social interaction and non-social behavior at different timings and in different directions. One
recent study observed that rats exposed to Pb made fewer ultrasonic vocalizations than did control rats at
PND 4 and 12 but not at PND 7 and 10 (Tartaglione et al.. 2020). The authors additionally did not
observe differences between exposed and control rats on the homing test, which evaluates olfactory
discrimination and social preference.
In summary, the body of evidence is suggestive of, but not sufficient to infer, a causal
relationship between Pb exposure and social cognition and behavior in children. The strongest
evidence supporting this causality determination comes from a novel case-control study in twins that
provides strong support for a positive association between dentine Pb levels and autism risk. There are a
number of recent prospective epidemiologic studies that provide supporting evidence of a positive
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association of increases in BLLs with reduced social cognition and increased autistic behaviors in
children, but the evidence is not entirely consistent and is limited by the potential for unmeasured
confounding by maternal age or the potential for reverse causality due to the timing of exposure in studies
examining blood Pb levels. Furthermore, the wide range of social cognition measures used in the
evaluated studies simultaneously adds dimensionality and complicates interpretation of the results. Only
one recent experimental animal study on Pb exposure and social cognition was available. This study,
combined with the toxicological evidence reviewed in the previous ISA suggests that Pb exposure may
influence social cognition and communication, though the direction of these effects is inconsistent. Thus,
while the limited experimental animal evidence provides some coherence with the epidemiologic
evidence, a number of uncertainties remain. The key evidence, as it relates to the causal framework, is
summarized in Table 3-8.
Table 3-8 Evidence that is suggestive of, but not sufficient to infer, a causal
relationship between Pb exposure and social cognition and
behavior in children
Rationale for
Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels
Associated with Effects0
Consistent evidence
from a few high-
quality epidemiologic
studies with relevant
blood, bone, and
tooth Pb levels
Greater difference in tooth Pb
levels among twins discordant
for ASD status than among
concordant twins.
Lower scores on test of social
cognition in a prospective study
in South Korea in association
with earlier childhood and
concurrent mean BLLs.
Arora et al. (2017)
Kimetal. (2016)
Deciduous tooth Pb NR
(early and postnatal Pb
levels)
Child blood Pb GM:
7-8 yr: 1.64 pg/dL
9-10 yr: 1.58 pg/dL
11—12 yr: 1.58 pg/dL
Evidence from multiple
prospective cohort studies for
small decrements in scores on
tests of social cognition among
children without autism ages
6.5 mo-3 yr at low levels of
exposure.
Shekhawat et al. (2021)
Rvaiel et al. (2021)
Cord blood Pb GM: 4.14
pg/dL
Mat. blood Pb GM (SD):
1st tri.: 5.27 (1.93) pg/dL
2nd tri.: 4.74 (1.96) pg/dL
3rd tri.: 4.98 (1.93) pg/dL
Infant blood GM (SD):
12 mo: 3.92 (1.80) pg/dL
24 mo: 3.49 (1.93) pg/dL
Zhou et al. (2017)
Mat. blood Pb GM (95%
CI): 3.30 (3.05, 3.57) pg/dL
Section 3.5.7.1
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RcL°u"alTtJ0r Key Evidence" References'
Determination3 Associated witn tracts
Epidemiologic
studies help rule out
chance, bias, and
confounding with
reasonable
confidence
Prospective studies had
population-based recruitment
with moderate to good follow-up
participation not conditional on
bone Pb/BLLs and social
cognition scores.
Section 3.5.7.1
All studies controlled for Table 3-13E
maternal age, education and/or
SES. Some controlled for
HOME score, maternal IQ, and
exposures to other pollutants.
Limited supporting
evidence from case-
control and cross-
sectional studies
Null findings from case-control
studies conducting adjusted
mean comparisons of ASD
cases and typically developing
controls with lower and higher
mean or median BLLs,
adjusting for maternal age,
various demographic and
lifestyle factors and dietary
consumption.
Skoaheim et al. (2021)
Rahbaret al. (2015)
Rahbaret al. (2021)
Mat. blood Pb GM
cases: 0.83 |jg/dL
controls: 0.88 |jg/dL
Child blood Pb GM (SD)
cases: 2.25 (2.23) |jg/dL
controls: 2.73 (1.85) pg/dL
Child Blood Pb GM
cases: 7.11 pg/dL
controls: 8.48 pg/dL
Greater BLLs among children
2-13 years old with
moderate/severe vs. mild
autism in a retrospective study.
Dong et al. (2022)
Child blood Pb mean (SD)
Mild: 2.58 (1.08) pg/dL
Moderate/severe: 3.25
(1.89) pg/dL
Null finding from cross-sectional
study. Lacked control for
maternal age at delivery.
Ruebner et al. (2019)
Child blood Pb med: 1.2
pg/dL
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RcL°u"alTtJ0r Key Evidence" References'
Determination3 Associated witn tracts
Limited experimental Mixed evidence of enhanced or
animal evidence at reduced social interaction
relevant exposures behavior among Pb-exposed
mice. Some suggestion of sex-
specific effect modification.
Reduced ultrasonic
vocalizations among Pb-
exposed rats. No evidence of a
difference on homing tests of
olfactory discrimination and
social preference.
Donald et al. (1986)
Donald et al. (1987)
Tartaqlione et al. (2020) Med blood Pb after
exposure during
pregnancy and lactation:
0.26 |jg/mL PND 23
ASD = autism spectrum disorder; BLL = blood lead level; CI = confidence interval; GM = geometric mean; HOME = Health
Outcomes and Measures of the Environment; IQ = intelligence quotient; Mat = maternal; med = median; mo = month(s); NR = not
reported; Pb = lead; PND = postnatal day; SD = standard deviation; SES = socioeconomic status; tri = trimester; yr = year(s).
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the
Preamble to the ISAs fU.S. EPA. 2015).
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and,
where applicable, to uncertainties or inconsistencies. References to earlier sections indicate where the full body of evidence is
described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
3.6 Nervous System Effects Ascertained during Adult Lifestages
The strongest evidence of Pb-associated nervous system effects in adults without occupational
exposure pertained to cumulative exposure and cognitive effects (U.S. EPA, 2013). Prospective studies
indicated associations of higher baseline tibia (means 19, 20 (ig/g) or patella (mean 25 |ig/g) Pb levels
with declines in cognitive function in adults (age >50 years) over 2- to 4-year periods. Pb-associated
cognitive function decrements were found with adjustment for potential confounding factors such as age,
education, SES, current alcohol use, and current smoking. Supporting evidence was provided by cross-
sectional studies, which found stronger associations with bone Pb level than concurrent BLL. Cross-
sectional studies also considered more potential confounding factors, including dietary factors, physical
activity, medication use, and comorbid conditions. The multiple exposures and health outcomes examined
in many studies reduced the likelihood of biased participation specifically by adults with higher Pb
exposure and lower cognitive function. Uncertainties remained due to residual confounding by age and
lack of information on the patterns of exposure associated with the BLLs observed in the epidemiologic
studies.
3.6.1 Cognitive Function in Adults
The evidence reviewed in the 2013 Pb ISA was sufficient to conclude that "a causal relationship
is likely to exist" between long-term cumulative Pb exposure and cognitive function decrements in adults
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(U.S. EPA, 2013). Prospective studies of the Normative Aging Study (NAS) and Baltimore Memory
Study (BMS) cohorts indicated associations of higher baseline tibia (means 19, 20 (ig/g) or patella (mean
25 (ig/g) Pb levels with declines in cognitive function in adults (age >50 years) over 2- to 4-year periods
among adults without occupational exposure (see Table 4-10 (U.S. EPA. 2013)). While the specific
covariates differed between studies, these bone Pb-associated cognitive function decrements were found
with adjustment for potential confounding factors such as age, education, SES, current alcohol use, and
current smoking. Supporting evidence was provided by cross-sectional analyses of the NAS, BMS, and
the Nurses" Health Study (NHS), which found stronger associations with bone Pb level than concurrent
BLL indicating the relative importance of long-term Pb exposure. Cross-sectional analyses considered
more potential confounding factors, including dietary factors, physical activity, medication use, and
comorbid conditions. The multiple exposures and health outcomes examined in many studies reduced the
likelihood of biased participation specifically by adults with higher Pb exposure and lower cognitive
function. The effects of recent Pb exposures on cognitive function decrements in adults were indicated in
Pb-exposed workers by associations found with BLLs, although these studies did not consider potential
confounding by other workplace exposures. The biological plausibility for the effects of Pb exposure on
cognitive function decrements in adults was provided by findings that relevant lifetime Pb exposures from
gestation, birth, or after weaning induce learning impairments in adult animals and by evidence for the
effects of Pb altering neurotransmitter function in the hippocampus, prefrontal cortex, and nucleus
accumbens (U.S. EPA, 2013).
Recent epidemiologic studies provide consistent evidence that higher cumulative exposure
indicated by bone Pb levels or childhood BLLs are associated with decrements in cognitive function
during young-, mid- or older-adulthood periods (Table 3-14E). Across populations, higher Pb levels were
associated with decrements in FSIQ, global cognitive function, executive function, visuospatial and
visuomotor skills, language, and memory. Much of this evidence was provided by extended analyses
(about 15 years of follow-up data) of the NAS and NHS cohorts considered in the 2013 Pb ISA, and
prospective cohort studies from Sweden and New Zealand that explored the effects of early childhood Pb
exposure (7-12 years) on IQ and various cognitive domains during young adulthood (18-19 years).
Findings from these recent prospective cohort studies, emphasize the important role of early childhood Pb
exposure and persistent effects on adult cognition after adjustments of various sociodemographic factors
and maternal and childhood IQ. Overall, the longitudinal design with longer follow-up periods, multiple
and repeatedly measured cognitive outcomes, and multiple risk factors and confounders accounted for in
the studies reduce the bias and strengthen the study findings related to the effects of Pb exposure on adult
IQ and cognitive function. Recent evidence from animal studies provide support that postnatal exposure
to Pb (either during adolescence or continuing into adulthood) negatively affects learning and memory in
rodents. Additionally, adult rodents exposed during early developmental periods displayed impairments in
tests of learning and memory conducted in adulthood (reviewed in Section 3.4). This suggests that early
life Pb exposure contributes to cognitive dysfunction that persists into adulthood, which is new evidence
in this review. Additionally, animals exposed to Pb during adulthood display similar cognitive
impairment, though there is still uncertainty regarding the influence of age on Pb exposures during
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adulthood. These studies add to the current evidence base suggesting a potential role of both early and
later life Pb exposures and biological plausibility for the effects of Pb exposure on cognitive function
decrements in adults.
A summary of the recent evidence, which is interpreted in the context of the entire body of
evidence, is provided in the subsequent sections. Measures of central tendency for Pb biomarker levels
used in each study, along with other study-specific details, including study population characteristics and
select effect estimates, are highlighted in Section 3.7, Table 3-14E (Epidemiology) and Table 3-4T
(Toxicology).
3.6.1.1 Epidemiologic Studies of Cognitive Function in Adults
Studies in the 2013 Pb ISA (U.S. EPA. 2013) found that higher bone Pb levels, indicating long-
term exposure to Pb, were associated with decrements in cognitive function in adults without
occupational Pb exposure. There was variability in associations across the various domains of cognitive
function tested within studies; however, higher bone Pb levels were associated with poorer performance
in most of the tests conducted. Further, discordant Pb associations across domains of cognitive function
are likely to reflect biologic variability or differences in the outcome pathophysiology. Across
populations, higher bone Pb levels were associated with decrements in executive function, visuospatial
skills, learning, and memory. Much of this evidence was provided by analyses of the BMS and NAS, with
additional findings reported in the NHS and smaller populations. The strongest evidence for bone Pb-
associated cognitive decrements demonstrated that higher tibia (means: 19, 20 (ig/g) and patella (mean:
25 (ig/g) bone Pb levels measured at baseline were associated with subsequent declines in cognitive
function over 2- to 4-year periods (Bandeen-Roche et al., 2009; Weisskopf et al., 2007). These findings
indicated that long-term Pb exposure may contribute to ongoing declines in cognitive function in adults.
These associations were found with adjustment for potential confounding by age, education, smoking, and
alcohol use in the NAS and age, sex, race, household wealth, and education in the BMS. While the NAS
and Nurses" Health Study included primarily white men and white women, respectively, the BMS
examined a more diverse population of men and women of various races and ethnicities.
In a recent analysis of the NAS cohort, Farooqui et al. (2017) examined the associations between
long-term Pb exposure quantified using bone biomarkers (mean patella Pb: 30.6 (ig/g and tibia Pb: 21.6
(ig/g) and longitudinal changes in cognition (repeatedly measured up to five visits over the 15 years
follow-up period) adjusted for age at the first cognitive test, education level, baseline smoking status, and
alcohol intake. The study found that higher patella bone Pb concentration (IQR: 21 |ig/g) was associated
with a 0.062 point lower baseline Mini Mental State Examination (MMSE) score (95% CI: -0.012,
0.003), 0.008 units/year MMSE decline (95% CI: -0.015, 0) over 15 years, and an increased risk of
having an MMSE score below 25 (threshold considered to represent cognitively not normal or at risk for
dementia) (hazard ratio [HR] = 1.10 (95% CI: 0.99, 1.21)). Similar but weaker and less precise
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associations were observed when tibia Pb and MMSE outcomes were assessed. The study also used
"global cognition" as a separate proxy for worsening cognitive impairment, and combined seven test
scores assessed in NES2, CERAD, and WAIS-R. Weaker associations were observed between both
patella and tibia Pb and global cognition (both baseline and longitudinal change). When separate
cognitive domains were assessed, patella Pb was associated with faster longitudinal decline in language
and memory domains, whereas similar but weaker associations were observed with tibia Pb.
A longitudinal study of women aged 45-74 enrolled in the NHS cohort (Power et al.. 2014) added
to the evidence provided by previous analyses. The authors examined the associations between Pb
exposures using bone and blood biomarkers (mean patella Pb: 12.6 (ig/g and tibia Pb: 10.5 |ig/g: mean
blood Pb: 2.9 (ig/dL) and cognitive decline (repeatedly measured using a telephone battery of cognitive
tests assessing learning, memory, executive function, and attention during 2-4 waves over the 13-year
follow-up period). Results were adjusted for alcohol consumption, smoking status, education, husband's
education, menopausal status and hormone therapy use, physical activity, ibuprofen use, aspirin use,
vitamin E supplementation, percentage of residential census tract of white race or ethnicity, and median
income of residential census track. A weak and imprecise association was observed for an excess annual
decline in the overall cognitive test Z-score per SD increase in tibia bone Pb concentration (-0.002
standard units; 95% CI: -0.005, 0.000). When individual cognitive tests were considered, a decline on the
East Boston Memory Test as well as immediate (a measure of episodic memory) and category fluency (a
measure of executive function and memory) was observed in relation to increased tibia Pb concentration
(Power et al.. 2014). There was little evidence for associations between patella Pb or blood Pb and the
decline in overall cognition, verbal memory, or individual cognitive tests.
Associations between either tibia or patella Pb concentration and cognitive function decline
observed in these longitudinal studies are supported by (Weuve et al.. 2013). a cross-sectional study that
assessed bone Pb and cognitive function (assessed using telephone cognitive assessment battery) among
participants from an existing case-control study of Parkinson's disease. The Pb concentration observed in
this study is similar to that reported in the NHS cohort. Separate analyses were performed for the group of
participants with PD and for all participants including both PD and control groups. Analysis of the PD
group showed that higher tibia Pb was significantly associated with worse overall performance (as shown
by the global cognitive score) and worse performance on the majority of telephone cognitive tests (in the
model adjusted for age at cognitive assessment, sex, race, education, smoking history). Patella Pb
concentration, however, was not consistently associated with cognitive performance. In the model with
both PD and control groups, interactions were observed for the association between tibia Pb and global
cognition by case-control status. Among the participants, a 10 (ig/g increase in tibia Pb corresponded to a
decrease in the global cognitive score by 0.12 standard units (95% CI: -0.22, -0.01), but the association
among controls was weak (0.06 standard units, 95% CI: -0.09, 0.20).
Evidence of the cognitive effects of cumulative Pb exposure observed in the NAS and NHS
studies is strengthened and extended further by the findings from cohort studies that examined the
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associations of childhood Pb exposure and continued long-term exposures with cognitive impairments in
young adults (18-19 years) (Skerfving et al.. 2015) or in mid-adulthood (38 or 45 years of age) (Reuben
et al.. 2020; Reuben et al.. 2017). SkerfVing et al. (2015) included samples of 7-12 year-old school
children in southern Sweden and followed them over time to examine the association between childhood
BLL (age 7-12 years old, mean blood Pb: 3.4 (ig/dL) and cognitive performance (IQ) assessed for
military conscription at 18-19 years of age using generalized linear models. The study found an IQ loss of
0.127 (95% CI: -0.209, -0.045) points per (ig/dL increase in childhood BLL for all participants and an IQ
loss of 0.204 (95% CI: -0.392, -0.016) points per (ig/dL increase in childhood BLL among those with
childhood BLLs <50 |ig/L. even in multivariable models adjusted for parent's income, education, and
father's IQ. Reuben et al. (2017) and Reuben et al. (2020) examined a New Zealand birth cohort with
participants born in 1972-1973 (a time when Pb exposure in New Zealand cities were higher than
international standards) who were part of the Dunedin Multidisciplinary Health and Development Study.
Infants were followed from birth through adulthood. Blood collection at 11 years of age provided blood
biomarker data for Pb (mean blood Pb: 10.99 (ig/dL). Cognitive performance was assessed using
objective tests of cognitive performance such as the WISC-R during childhood at ages 7 and 9 years.
Cognitive performance was also assessed using the WAIS-IV when participants were 38 years old
(Reuben et al.. 2017) and again when they were 45 years old (Reuben et al.. 2020). Reuben et al. (2020).
in addition to the objective tests, also included subjective reports of everyday cognitive functioning
(memory or attention problems) at age 45 years as provided by study participants and their nominated
informants. The studies examined the association between childhood blood Pb and adult cognitive
outcomes or cognitive decline (change in IQ score between childhood and mid-adulthood), using OLS
multiple regression models. Reuben et al. (2017) found that each 1 (ig/dL higher level of blood Pb in
childhood was associated with a 0.39-point lower score in adult FSIQ (95% CI: -0.67, -0.12), and a 0.32-
point decline (95% CI: -0.50, -0.15) after adjusting for sex, childhood IQ, maternal IQ, and childhood
SES. Similarly, with additional years of data (Reuben et al.. 2020) continued to show significant
associations between childhood BLL and IQ at 45 years of age. Each 1 (ig/dL higher childhood BLL was
associated with a -0.41 (95% CI: -0.68, -0.15) point decline in full-scale IQ from baseline. When using a
residualized change model to adjust for autocorrelation between baseline and follow-up IQ, the decline in
IQ was similar (-0.39 [95% CI: -0.58, -0.21]). The study also found that the relationship between
childhood blood Pb and adult IQ persisted and remained significant even after adjustment for brain
structure measures. Results pertaining to brain structure are discussed in Section 3.4.1.
Several recent cross-sectional studies also examined the associations of Pb exposure and
decrements in cognitive function. The majority used concurrent blood biomarkers (Sasaki and Carpenter.
2022; Xiao et al.. 2021; Przvbvla et al.. 2017; Souza-Talarico et al.. 2017; Khalil et al.. 2014; van
Wiingaarden et al.. 2011). A few studies used other biomarkers such as toenails (Meramat et al.. 2017).
bone (Weuve et al.. 2013). or urine (Sasaki and Carpenter. 2022). As described further below, results
were not entirely consistent across studies.
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Three of the cross-sectional studies examining the blood Pb-cognitive function association used
data from various NHANES cycles including participants aged 60-84 years (Sasaki and Carpenter. 2022;
Przvbvla et al.. 2017; van Wiingaarden et al.. 2011). van Wiingaarden et al. (2011) examined the
associations of blood Pb (mean blood Pb: 2.46 (ig/dL) with self-reported confusion and memory problems
using data from NHANES (1999-2008). Data from the 1999-2002 NHANES cycle were used to estimate
the association between BLL and performance on the Digit Symbol Substitution Test (DSST) scores.
After adjustment for age, sex, education level, ethnicity, poverty-income ratio (PIR), self-reported health
status, and comorbid conditions, no association of BLLs with self-reported confusion or memory
problems or DSST performance was observed (van Wiingaarden et al.. 2011). Two other studies,
Przvbvla et al. (2017) and Sasaki and Carpenter (2022). included the NHANES participants from 1999-
2002 and 2011-2014 cycles, respectively. These studies explored the cross-sectional associations of blood
and urine biomarkers of multiple metals and metalloids (separately and jointly) on cognitive function.
(Przvbvla et al.. 2017) used a path analysis approach to model multiple exposures of 14 chemicals
simultaneously while adjusting for multiple comparisons. The study found that the association of BLL
(Geometric mean: 2.17 (ig/dL) with lower cognition scores was attenuated when the model controlled for
smoking status. Specifically, a 1-SD increase in BLL was weakly associated with a slightly lower Digit
Symbol Coding (DSC) test score from WAIS-III (|3 = -0.10, 95% CI: -0.20, -0.00) after controlling for
co-exposure and sociodemographic covariates. The study also performed stratified analysis by sex and
age (above and below median age) and found a greater magnitude of associations for female and higher
age categories (>10%), despite a lack of statistical evidence of an interaction. (Sasaki and Carpenter.
2022) used two stage linear regression models. First, they performed single metal analyses separately for
each of seven metals or metalloids followed by a second analysis including multiple metals or metalloids
from the stage 1 analysis to examine the associations with immediate, delayed, and working memory
quantified using CERAD and DSST. When single metals were assessed, increased blood Pb concentration
was associated with decrements in performance on all three cognitive tests after adjusting for
sociodemographic, behavior, and clinical characteristics (immediate recall: |3 = -0.58, 95% CI: -0.91,
-0.24; delayed recall: |3 = -0.19, 95% CI: -0.35, -0.02; Digital Symbol Substitution: |3 = -1.08, 95% CI:
-2.12, -0.05). Multi-metal analysis stratified by age group (60-70 and >70 years old) suggested greater
declines in immediate recall among participants over the age of 70. Khalil et al. (2014) examined the
association between concurrent blood Pb concentration (mean blood Pb: 2.25 (ig/dL) and cognitive
function among a subset of non-institutionalized community dwelling non-Hispanic Caucasian men 65
years and older who participated in the Osteoporotic Fractures in Men Study (MrOS) cohort study.
Cognitive function was assessed using the Modified MMSE (3MS) and the Trail Making Test Part B.
Higher scores on the 3MS and faster time on the Trail Making Test Part B both represent better
performance. Multivariable analysis found no association between blood Pb concentration and cognitive
function (Khalil et al.. 2014).
Souza-Talarico et al. (2017) examined the association between blood Pb (and interactions
between blood Pb and Cd) and working memory capacity (WMC) in a population of 125 older adults
aged 50-82 years, in the metropolitan area of Sao Paulo, Brazil. The study also explored the mediating
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role of antioxidant capacity (using various oxidative stress biomarkers) in the heavy metals-memory
associations. Using regression models accounting for age, sex, income, and hemoglobin, the study did not
find an association between blood Pb (mean 2.1 (ig/dL) and WMC (|3 = 0.106, 95% CI:-0.208, 0.417);
however, an interaction between blood Pb and blood Cd level was observed as well as a significant
inverse association between the blood Cd x blood Pb interaction term and WMC was observed (Souza-
Talarico et al.. 2017). The Monte Carlo Method for Assessing Mediation test for mediation revealed that
the association between the blood Cd x blood Pb interaction term and WMC was significantly mediated
by total antioxidant capacity.
(Xiao et al.. 2021) examined the association between multiple metals (22 metals including Pb;
mean blood Pb: 5.15 (ig/dL) and cognitive function measured using MMSE in participants aged >60 years
from Guangxi, Southern China. The study used least absolute shrinkage and selection operator (LASSO)
penalized regression to identify main metals associated with cognitive function. Twelve metals (including
Pb) selected from LASSO were then explored in a multi-metal generalized linear regression model
adjusted for age, gender, education attainment, annual income, BMI, smoking, alcohol, insomnia, and
physical activity. No association was observed for blood Pb and cognitive function after adjustment for
other metals.
Notably, a limitation of cross-sectional studies of concurrent BLLs is that the relative contribution
of the recent versus past Pb exposure is not well characterized. A recent prospective study was designed
to address the uncertainties related to the exposure patterns associated with BLLs observed in studies of
adults (Yu et al.. 2021). Yu et al. (2021) examined the association of BLLs and neurocognitive
performance among newly hired employees at battery manufacturing and Pb recycling plants with no
previous occupational Pb exposure, a subset of participants in the Study for Promotion of Health in
Recycling Lead (SPHERL) cohort study. Baseline blood Pb concentration was measured, and the
participants were followed annually over a 2-year period to measure blood Pb biomarkers and assess if
higher recent occupational exposure to Pb was associated with neurocognitive dysfunction. The
participants completed the DSST and SCWT at baseline and annual follow-up visits. The geometric mean
blood Pb at baseline and first and second follow-up visits were 3.97 (ig/dL, 13.4 (ig/dL, and 12.8 (ig/dL,
respectively, showing an almost three-fold increase in blood Pb over the 2 years of occupational
exposure. The study used a linear mixed model to examine the changes in DSST and SCWT
corresponding to changes in blood Pb separately for the 1- and 2-year visits. Despite the three-fold
increase in blood Pb concentration, the study found no association between blood Pb and cognitive
function. The change in latency time and error rate based on the DSST test showed an increase from
baseline to follow-up, with an increase in the follow-up-to-baseline blood Pb concentration ratio, but the
association was weak and imprecise in the fully adjusted models (change in latency: 0.55%, 95% CI:
-0.33, 1.42; error rate: OR: 1.01, 95% CI: 1.00, 1.03)
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3.6.1.1.1 Summary
Longitudinal cohort studies evaluated in the 2013 Pb ISA found consistent evidence of an
association between increased long-term exposure to lead indicated by bone Pb levels and decreased
cognition in adults. Recent prospective cohort studies add to the body of evidence informing the
relationship between Pb exposure and cognitive performance in adults without occupational Pb exposure.
More specifically, recent cohort studies indicated that higher adult bone Pb levels, which indicate
cumulative Pb exposure (tibia mean range: 10.5, 21.6 (ig/g, patella mean range: 12.6, 30.6 |ig/g) or
childhood BLLs (mean range: 3.4 (ig/dL, 10.99 (ig/dL at 7-12 years of age), were associated with
decrements in cognitive function or IQ during young-, mid-, or older-adulthood periods (Table 3-14E).
There was some variability in the associations with various domains of cognitive function tested within
studies; however, variability in the associations observed across domains of cognitive function generally
reflects biologic variability or differences in the outcome pathophysiology rather than inconsistent study
results. Across studies, higher Pb levels were associated with decrements in FSIQ, global cognitive
function, executive function, visuospatial and visuomotor skills, language, and memory. Extended
analyses of the NAS and NHS cohorts with 13 to 15 years of follow-up add to the evidence base
(Farooaui et al.. 2017; Power et al.. 2014). These studies found associations of cumulative Pb exposure
with decrements in cognitive function in adults after adjustment for potential confounding by
combinations of factors including demographic, socioeconomic, behavioral, clinical, and neighborhood-
level factors. In addition, findings from recent prospective cohort studies in Sweden and New Zealand
that explored the effects of Pb exposure during childhood lifestages (7-12 years) on IQ and cognitive
effects during young adulthood (18-19 years) (Skcrfving et al.. 2015) and mid-adulthood (38-45 years)
(Reuben et al.. 2020; Reuben et al.. 2017). These studies found that higher childhood BLLs were
associated with declines in IQ ascertained in adulthood after adjustment for demographic and
socioeconomic factors, maternal IQ, and childhood IQ scores. These findings provide new insight into the
persistence of Pb-associated cognitive function decrements. Overall, the longitudinal design with longer
follow-up periods, multiple and repeatedly measured cognitive outcomes, and multiple risk factors and
confounders accounted for in epidemiologic studies investigating long-term cumulative exposure and
early childhood exposure reduce uncertainties and strengthen the evidence related to the association of Pb
exposure with cognitive function in adulthood. Sex (male versus female, premenopause versus
postmenopause) and age (young versus mid-aged versus old-aged adults) differences in bone kinetics and
turnover, as well as disease comorbidity, particularly at middle- and older-adulthood lifestages may
potentially lead to differences in bone Pb and blood Pb levels and add complexity when modeling the
associations since inclusion of only age or sex in the model may not fully account for these differences.
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3.6.1.2
Toxicological Studies of Cognitive Function in Adults
3.6.1.2.1 Learning and Memory - Morris Water Maze
This section specifically reviews studies that exposed animals to Pb during either adulthood or
late adolescence. Studies that exposed animals during development (i.e., pregestation, gestation, lactation)
are reviewed in Section 3.5.1.3.2. Animals exposed to Pb via drinking water in adulthood displayed
impaired learning and memory. Using the Morris water maze, Mansouri et al. (2012) found that short-
term Pb exposure (50 mg/L Pb in drinking water, PND 70 to 100), which produced mean BLLs of 8
(ig/dL, significantly impaired both learning and memory, though the magnitude of the effect on memory
was smaller compared with the effect observed in other studies of developmental exposure (Section
3.5.1.3.2). Also using the Morris water maze, Mansouri et al. (2013) reported that long-term Pb exposure
(50 ppm in drinking water, PND 60 to 240), which produced peak BLLs of 11-19 (ig/dL, significantly
impaired learning and memory performance in both sexes. Cognitive impairment was also observed in
studies that utilized daily administration of Pb via gavage. Singh et al. (2019) reported significantly
increased escape latencies and path lengths (i.e., distance traveled to reach the platform, another measure
of learning) in exposed rats following long-term Pb exposure via gavage (2.5 mg/kg, PND 90 to 180),
which produced peak BLLs of 28 (ig/dL. No probe phase was conducted in this study. Additionally, Su et
al. (2016) found that male rats gavaged with Pb solutions daily (200 ppm, PND 20 to 76) displayed
significant impairments in both the learning and memory components of the Morris water maze.
In two recent studies, Zou et al. (2015) and Han et al. (2014) reported significant learning and
memory deficits following short-term exposure of juvenile animals. Zou et al. (2015) exposed mice from
PND 35 to 56 with mean BLLs of 22 (ig/dL, while Han et al. (2014) exposed rats from PND 21 to 42 and
reported mean BLLs of 15 (ig/dL. Other studies examined the effects of long-term Pb exposure on
juvenile rodents and found similar effects. For example, (An et al.. 2014) exposed groups of juvenile rats
to multiple doses of Pb for 56 days (PND 28 to 84). All examined doses (100, 200, and 300 ppm in
drinking water) produced BLLs relevant to this ISA. At the time of Morris water maze assessment (PND
84), the mean BLL ranged from 11 to 23 (ig/dL. All exposed animals displayed impaired memory during
the probe trial relative to controls. Only animals in the two highest dose groups were reported to show
learning deficits during training, suggesting that, with juvenile exposures, Pb may have a greater effect on
memory than learning processes. Another study that exposed juvenile mice to Pb in drinking water (0.2%)
for 90 days (PND 28 to 112) assessed Morris water maze performance in the same animals at multiple
time points during and immediately following exposure (Wu et al.. 2020b). This long-term exposure
produced relatively high mean BLLs of 28 (ig/dL, and all exposed animals showed signs of impaired
learning and memory in the maze. Interestingly, both measures of cognition improved in the exposed
animals over time, which may reflect either increasing familiarity with the task or clearance of Pb over
time. In contrast to all other studies in young and juvenile animals, Li et al. (2013) reported that rats given
Pb in drinking water for 84 days (from PND 28 to 112) with peak BLLs of 16 (ig/dL showed no
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indication of learning or memory impairment in the Morris water maze. Despite this one discrepant study,
recent evidence supports the notion that postnatal exposure to Pb (either during adolescence or continuing
into adulthood) negatively affects learning and memory in rodents, which contrasts with several of the
key studies reviewed in the previous ISA.
3.6.1.2.2 Summary
Four recent studies of rodents with exposure resulting in mean BLLs <30 (ig/dL add to the
evidence informing the association of both short- and long-term Pb exposure during adulthood with
measures of learning and memory in rodents. While these studies are consistent with one another,
toxicological evidence for the effects of Pb on cognitive function in adults remains limited. Additionally,
a few recent studies in juvenile rodents also provide some support for the association between postnatal
Pb exposure either during adolescence or continuing into adulthood and cognitive impairment,
specifically learning and memory.
3.6.1.3 Relevant Issues for Interpreting the Evidence Base
3.6.1.3.1 Concentration-Response Function
The 2013 Pb ISA reviewed a small number of studies that examined the shape of the C-R
relationship between blood or bone Pb levels and cognitive function. Studies using BMS and NAS
cohorts assessed nonlinearity using quadratic terms, penalized splines, or visual inspection of bivariate
plots. Prospective analyses of the NAS cohorts provided some evidence of nonlinearity (Wang ct al..
2007; Weisskopf et al.. 2007) Figures 4-7 and 4-8 from 2013 Pb ISA). Weisskopf et al. (2007) found that
a 20 (ig/g difference in patella Pb level was associated with a 0.07-ms increase in response latency (95%
CI: 0.04, 0.12; larger values mean slower reaction times in the pattern comparison test) among all men
and a 0.15-ms increase among men with patella Pb level <60 |ig/g. These results suggest that Pb-
associated latency worsens with increasing Pb up to 60 (ig/g and levels off at higher values. Wang et al.
(2007) found that among NAS men with an HFE gene variant, there was a larger decline in MMSE score
(a global examination of cognitive function with low scores indicating poor cognitive performance) per
unit increase in tibia Pb level at higher tibia Pb levels.
In the current review, the shape of the C-R function was not assessed in studies that examined the
associations of Pb biomarkers with cognitive function in adults. The majority of studies selected
analytical models that assumed linear associations in the Pb-cognitive function associations. A few
studies in the recent review examining the influence of childhood Pb exposure on cognitive impairments
at the young- (18-19 years) (SkerfVing et al.. 2015) or mid-adulthood periods (38 or 45 years of age).
Reuben et al. (2017) and Reuben et al. (2020) performed separate analyses for the subsets of the
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population exposed to higher and lower Pb levels to examine possible nonlinear relationships or threshold
effects. Reuben et al. (2017) found a 1.97-IQ-point reduction in adulthood (95% CI: -3.34, -0.59) for the
overall sample, a 4.25-IQ-point reduction for individuals above the level of concern, and a 2.73-IQ-point
reduction for individuals below the level of concern for each 5-(.ig/dL increase in the childhood blood Pb.
Similarly, IQ decline from childhood to adulthood suggested a mean decline of a 1.61 IQ points (95% CI:
-2.48, -0.74) in adulthood for the overall sample, a mean decline of 1.68 IQ points for participants above
the level of concern, and a mean increase of 1.22 IQ points for participants below the level of concern for
each 5-(ig/dL increase in the childhood blood Pb. Similar results were observed by (Reuben et al.. 2020)
in the same population studied in (Reuben et al.. 2017) and followed till 45 years of age. SkerfVing et al.
(2015) examined the influence of early childhood Pb exposure on long-term cognitive impairments at
young adulthood (18-19 years) using generalized linear models. The study found an IQ loss of 0.127
(95% CI: -0.209, -0.045) points per (ig/dL increase in childhood BLLs for all participants, and a slightly
larger IQ loss (i.e., 0.204 [95% CI: -0.392, -0.016] point per (ig/dL increase in childhood BLL) for the
populations with childhood BLLs <50 |ig/L. Reuben et al. (2017) and Reuben et al. (2020) examined the
association of childhood blood Pb (11 years) with cognitive performance during mid-adulthood and
cognitive decline (change in IQ score between childhood and mid-adulthood) for the overall sample as
well as separately for participants above or below the historic level of concern (i.e., >10 (.ig/dL). Overall,
these childhood exposure studies suggested persistence and continued cognitive effects of childhood Pb
exposure through mid-adulthood, and the strength of associations were higher in magnitude for the
participants with childhood exposure above the historic level of concern (>10 (.ig/dL).
The limited recent toxicological evidence generally supports the dose-dependent effects of Pb on
cognitive function at relevant BLLs in adult animals. Only one study examined juvenile animals exposed
to multiple concentrations of Pb and reported greater decrements in learning and memory at higher doses
(An et al.. 2014).
3.6.1.3.2 Potentially At-Risk Populations
Age and Sex:
In the 2013 Pb ISA, an analysis using the NAS cohort reported an interaction between Pb and age
(Wright et al.. 2003). The study reported that the inverse association between age and cognitive function
was greater among those with high blood or patella Pb levels. Specifically, in the highest quartile of
patella Pb, each year increase in age led to a four-fold steeper decline in the MMSE score relative to the
effect of age in the lowest quartile of patella Pb. Effect estimates were in the same direction for tibia Pb,
but the interaction was not statistically significant.
Two recent epidemiologic studies using NHANES data from the 1999-2002 and 2011-2014
cycles explored the cross-sectional associations of blood and urine biomarkers of multiple metals and
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metalloids (separately and jointly) with cognitive function and provided some insights into potential
effect modifications of Pb-associated decrements in cognitive function (Sasaki and Carpenter. 2022;
Przvbvla et al.. 2017). Stratified analysis by sex and/or age groups (above and below median age groups)
performed in these studies suggested a greater magnitude of Pb-cognitive function associations (beta
estimates >10%) for females and for older age categories, however, the statistical test for the interactions
suggested no difference between the sex and age categories.
Toxicological studies investigating potential sex differences in Pb-induced cognitive impairment
are limited. A study by Mansouri et al. (2013) reported that Pb produced similar decrements in learning
and memory in both male and female animals. Given the lack of toxicological evidence available, the
possible influence of sex on Pb-induced cognitive impairment in adult animals remains unclear.
Pre-existing conditions:
One study evaluated the association of bone Pb (mean ranges for various age groups: tibia Pb:
4.4-9.2 (ig/g; patella Pb: 5.9-15.2 |ig/g) with cognitive function among individuals with PD and controls
participating in a case-control study (Weuve et al.. 2013). The patella Pb and tibia Pb concentrations
reported in this study for all study participants increased with increasing age. The highest Pb
concentrations were found in study participants in the 75-81 years old category, and the lowest
concentrations were found in participants in the 54-65 years old category. When the data were analyzed
separately for participants with PD, higher tibia Pb concentration was significantly associated with lower
scores on all of the telephone cognitive tests (adjusted difference in scores per 10 (ig/g increase in bone
Pb: Telephone Interview for Cognitive status (TICS) test: -0.20 [-0.4, -0.00]; digit span forward: -0.23
[-0.43, -0.03]; digit span backward: -0.19 [-0.37, -0.00]) and global cognitive score (adjusted
difference in scores per 10 (ig/g increase in bone Pb: -0.13 [-0.25, -0.01]). When the overall (cases and
control) data were analyzed, significant interactions were observed for the association between tibia Pb
and global cognition by case-control status. Participants with PD showed worse scores compared with
controls (1 SD increase in tibia Pb led to worsening of the global cognitive score by 0.12 units [95% CI:
-0.22, -0.01] among cases; controls: 0.06 [95% CI: -0.09, 0.20]).
Genetics:
Studies investigating the association between Pb levels and cognitive function in 2013 Pb ISA
extensively evaluated the effect modification by ALAD and HFE gene variants. The evidence was
provided by an NHANES analysis (Krieg et al.. 2009) as well as multiple analyses from the NAS cohort
examining different tests of cognitive function (Raian et al.. 2008; Weuve et al.. 2006). In the study using
a cohort from NHANES III, associations with concurrent BLLs were more pronounced in groups with CC
and CG ALAD genotypes (i.e., ALAD2 carriers) for several indices of cognitive function (Krieg et al..
2009). In the NAS cohort of men, Weuve et al. (2006) found that higher concurrent BLL but not bone Pb
level was associated with a larger decrease in a test of general cognitive function among ALAD2 carriers.
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Another NAS study examined the function of specific cognitive domains (e.g., vocabulary, memory,
visuospatial skills) and found variable evidence for effect modification by ALAD genotype across tests
(Raian et al.. 2008). For example, among ALAD2 carriers, concurrent BLL was associated with a more
pronounced decrease in vocabulary score but less pronounced decrease in a memory index and no
difference in the associations with other cognitive tests. For tibia and patella Pb levels, ALAD genotype
was found to modify associations with different tests, for example, executive function and perceptual
speed. It is not clear why the direction of effect modification would vary among different cognitive
domains. The limited number of populations examined, and the different cognitive tests performed in
each study, make it difficult to conclusively summarize findings for effect modification by ALAD
variants. However, in the limited available body of evidence, blood and bone Pb levels were generally
associated with lower cognitive function in ALAD2 carriers.
Longitudinal analysis of the NAS cohort also indicated that HFE gene variants modified the
blood Pb-cognition association (Wang et al.. 2007). Wang et al. (2007) found an IQR higher tibia Pb level
(15 (ig/g) was associated with a 0.22 point steeper annual decline (95% CI: -0.39, -0.05) in the MMSE,
which assesses cognitive impairment in a number of domains, among the men with at least one HFE
variant allele (H63D or C282Y variant). The association was found to be nonlinear, with larger Pb-
associated declines observed at higher tibia Pb levels. Tibia Pb level was not associated with a decline in
MMSE score in men with the HFE wildtype genotype. Moreover, the deleterious association between
tibia Pb and cognitive decline appeared progressively worse in participants with increasingly more copies
of HFE variant alleles (p-trend = 0.008). These findings suggest that HFE polymorphisms greatly enhance
susceptibility to Pb-related cognitive impairment in a pattern consistent with allelic dose.
None of the studies in the current review examined the effect modification by genetic variants.
Other Metal Exposure:
Various studies that examined other metals either evaluated the relationship of each metal
separately with the outcomes of interest or included multiple metals jointly in the model (Sasaki and
Carpenter. 2022; Xiao et al.. 2021; Przvbvla et al.. 2017). In a study of adults 50 to 82 years old in Sao
Paulo, Brazil, Souza-Talarico et al. (2017) examined the associations of heavy metals (Cd and Pb) in
blood and WMC separately as well as together in a model for metal interactions. The study found no
significant association between blood Pb and WMC in the model including Pb only, but significant
interactions were observed between blood Cd and blood Pb and the inverse association with WMC (|3 =
-0.38, p< 0.001).
3.6.1.3.3 Lifestages
The identification of critical lifestages and time periods of Pb exposure is complicated by the fact
that the majority of adult cognitive studies used concurrent adult BLLs. Although possibly affected by
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recent exposure, BLLs are also influenced by Pb stored in bone. Thus, associations in adult studies using
concurrent BLL may reflect the effects of past and recent Pb exposures on cognitive outcomes. Some
cohort studies in the 2013 Pb ISA and the current review using bone Pb suggested the effects of
cumulative long-term Pb exposure on cognitive impairment during adulthood. However, it is still difficult
to specifically identify exposures at particular lifestages (prenatal, infancy, early and late childhood, early
adulthood, etc.) that could have led to the long-term cognitive impairment observed in these studies. Few
recent prospective studies evaluating early childhood exposures at 7-12 years of age and long-term
cognitive impairment and decline at young- (18-19 years) (Skerfving et al.. 2015) and mid-adulthood (38
or 45 years of age) (Reuben et al.. 2020; Reuben et al.. 2017) provided an insight into critical lifestages
(i.e., early childhood and persistence of the cognitive effects through adulthood). Skerfving et al. (2015)
examined the association of childhood BLL in children from southern Sweden (age 7-12 years old, mean
blood Pb: 3.4 (ig/dL) with cognitive performance (IQ) at the age of 18-19 years. They found an IQ loss of
0.127 (-0.209, -0.045) points per (ig/dL increase in childhood BLL for all participants and an IQ loss of
0.204 (-0.392, -0.016) points per (ig/dL increase in childhood BLL among those with childhood BLLs
<50 |ig/L even in multivariable models adjusted for parent's income, education, and father's IQ. Reuben
et al. (2017) and Reuben et al. (2020) followed a New Zealand birth cohort to examine the association of
childhood Pb level (age 11 years, mean blood Pb: 10.99 (ig/dL) with cognitive performance and decline at
38 years (Reuben et al.. 2017) and 45 years (Reuben et al.. 2020). Reuben et al. (2017) found that each 5-
(ig/dL higher level of blood Pb in childhood was associated with a 1.97-point decrease in IQ score (95%
CI: -3.34, -0.59) and a 1.61-point decline (95% CI: -2.48, -0.74) in adult FSIQ after adjusting for sex,
childhood IQ, maternal IQ, and childhood SES. With additional years of data, Reuben et al. (2020) also
showed a significant association between childhood BLL and IQ at 45 years of age. Each 5-(ig/dL higher
level of blood Pb in childhood was associated with a 2.07-point decrease in the full-IQ score (95% CI:
-3.39, -0.74), and a 1.97-point decline (95% CI: -2.92, -1.03) after adjusting for covariates. These study
findings suggest that Pb exposure during childhood lifestages can influence cognition in adulthood.
Overall, recent rodent studies of learning and memory evaluating Pb exposure at various
lifestages suggest cognitive impairment. However, the magnitude of the effect at different lifestages has
been shown to differ. Adult animals may be less sensitive than juvenile animals, and juvenile animals
may be less sensitive than animals exposed during development (reviewed in Section 3.5.1). This general
pattern is consistent with evidence describing critical windows for brain development (Section 3.3).
Additionally, critical evidence for the association of Pb with cognitive impairment across lifestages comes
from a series of studies describing the effects of lifetime Pb exposure on nonhuman primates (Rice. 1992;
Rice and Gilbert. 1990a; Rice. 1990; Rice and Karpinski. 1988). Cynomolgus monkeys (Macaco
fascicularis) were dosed continuously from birth and tested repeatedly throughout their lifetime. While
these exposures yielded BLLs beyond values considered relevant for the current assessment (>30 (.ig/dL).
they provide key evidence of Pb-induced cognitive impairments that persisted into adulthood in a
translationally relevant species. However, given the limited number of studies conducted in juvenile and
adult animals, and the lack of studies examining the same endpoint across multiple age groups, the precise
role of exposures at various stages on the cognitive effects in adult animals remains unclear.
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3.6.1.4
Summary and Causality Determination: Cognitive Function in Adults
The 2013 Pb ISA (U.S. EPA, 2013) concluded that the available evidence was sufficient to
conclude "a causal relationship is likely to exist" between long-term cumulative Pb exposure and
cognitive function decrements in adults. This causality determination was based on a small body of
prospective studies that indicated strong associations of higher baseline tibia (means 19, 20 (ig/g) or
patella (mean 25 (ig/g) Pb levels with declines in cognitive function in adults (age >50 years) over 2- to 4-
year periods among adults without occupational exposure (i.e., NAS and BMS cohorts). Supporting
evidence was provided by analyses of the NAS, BMS, and NHS cohorts, that found stronger associations
with cognitive impairment for cumulative exposure (i.e., bone Pb level) than for concurrent BLL, which
reflects recent Pb exposure and Pb that has been mobilized from the bone. The timing, frequency,
duration, and magnitude of Pb exposures that contributed to the associations observed with blood Pb
levels were not discernable from cross-sectional associations reported in these studies. The biological
plausibility for the effects of Pb exposure on cognitive function decrements in adults was provided by
findings that relevant lifetime Pb exposures from gestation, birth, or after weaning induce learning
impairments in adult animals and by evidence for the effects of Pb altering neurotransmitter function in
the hippocampus, prefrontal cortex, and nucleus accumbens.
Results from recent epidemiologic and animal studies add to the evidence base reviewed in the
2013 Pb ISA. Recent epidemiologic studies consistently report that higher cumulative Pb exposure (i.e.,
bone Pb levels) or childhood BLLs, were associated with poor cognitive performance or decrements in
cognitive function during young-, mid-, or older-adulthood periods (Table 3-14E). Across populations,
higher Pb levels were associated with decrements in FSIQ, global cognitive function, executive function,
visuospatial and visuomotor skills, language, and memory. Discordant Pb associations across domains of
cognitive function are likely to reflect inherent biologic variability or differences in the outcome
pathophysiology as opposed to inconsistency in the evidence. Much of this evidence on adult cognitive
outcomes was obtained from analyses of the NAS and NHS cohorts, including recent analyses that
extended follow-up periods beyond the analyses evaluated in the 2013 Pb ISA. Recent evidence also
comes from early childhood exposure cohort studies conducted in Sweden and New Zealand. These
studies strengthen findings that childhood Pb exposures are associated with decrements in IQ and
cognitive function during young- and mid-adulthood. Longitudinal study designs with longer follow-up
periods, multiple and repeatedly measured cognitive outcomes, and multiple risk factors and confounders
accounted for in the studies reduce the bias and strengthen the study findings related to Pb exposure and
adult cognitive function. Further, significant findings from new studies specifically investigating the
influence of early childhood Pb exposure on adult IQ and cognitive outcomes, even after adjustments for
various confounders including childhood IQ, provide evidence for the role of early childhood Pb
exposures on decrements in cognitive function in adulthood.
Strong evidence for cognitive function declines associated with cumulative Pb exposures was
provided by prospective cohort studies that demonstrated increased bone Pb levels (tibia mean: 10.5, 21.6
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(ig/g, patella mean: 12.6, 30.6 |ig/g) measured at baseline were associated with cognitive decline over the
follow-up period of 13-15 years (Farooqui et al.. 2017; Power et al.. 2014). Findings from these studies
suggest that long-term Pb exposure may contribute to ongoing declines in cognitive function in adults.
These associations remained significant even after adjustment for potential confounding by combinations
of factors including demographic, socioeconomic, behavioral, clinical, and neighborhood level factors.
Increased bone Pb level (tibia mean range: 4.4-9.2 (ig/g) was associated with cognitive function outcomes
among cases of PD (Wcuvc et al.. 2013). Additional support for the effects of cumulative or past Pb
exposure is provided by an analysis of past blood Pb exposures during childhood (either low or high Pb
exposure scenarios; blood Pb mean: 3.4 (ig/dL at 7-12 years, 10.99 (ig/dL at 11 years) and studies that
followed study participants through young or mid-adulthood (Reuben et al.. 2020; Reuben et al.. 2017;
SkerfVing et al.. 2015). These studies indicated that higher childhood BLL was associated with declines in
IQ at 18 to 19 years old and at 38 years or 45 years old.
Findings from cross-sectional studies that assessed the relationships of concurrent blood (2.1
(ig/dL to 5.1 (ig/dL) and cognitive function outcomes were more mixed. Concurrent blood Pb level does
not clearly indicate recent Pb exposure in adults because Pb is mobilized from the bone in various adult
lifestages complicating the interpretation of these studies. Two studies including NHANES data found
inverse association between concurrent BLLs and cognitive outcomes (Sasaki and Carpenter. 2022;
Przvbvla et al.. 2017). while others suggested null associations (Xiao et al.. 2021; Souza-Talarico et al..
2017; Khalil et al.. 2014; van Wijngaardcn et al.. 2011). The NHANES studies demonstrating significant
associations considered multiple metals in their analytical models including Pb, used advanced model
approaches to handle multiple exposures and issues around multiple comparison and multi-collinearity,
and adjusted for sociodemographic, behavioral, and clinical characteristics. These approaches reduced the
bias and uncertainty in the study findings. A recent study using a prospective design addressed some of
the concern around the health effects of recent Pb exposure (Yu et al.. 2021). The study included a group
of individuals with no prior occupational exposure and recently hired young workers at battery
manufacturing and Pb recycling plants. The association between neurocognitive performance and blood
Pb was examined prior to and up to 2 years after the first occupational exposure (geometric mean
baseline: 3.97 (ig/dL; 13.4 (ig/dL, and 12.8 (ig/dL at the first and second follow-up visits). The study did
not observe significant associations of changes in neurocognitive function in the workers with an over
three-fold increase in blood Pb concentration over the 2-year follow-up period, though the follow-up time
in this study may not have been adequate to detect the long-term effects of Pb on cognitive function.
Sex and age differences in bone kinetics and turnover could have contributed to differences in the
magnitude of associations observed for different bone Pb biomarkers and cognitive function in the NHS
and NAS cohorts. Specifically, the role of specific bone biomarkers (i.e., tibia or patella Pb) as indicators
of Pb exposure for specific age and sex groups in relation to individual cognitive domains is yet to be
fully understood. For instance, the findings from the NHS cohort with shorter follow-up, Weuve et al.
(2009). suggested that higher tibia Pb in women was inversely associated with the overall cognitive score.
The associations with the majority of the domain-specific cognitive scores were also negative (except for
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letter fluency, which was positive) but the estimates were imprecise. The recent extended analysis of the
NHS cohort Power et al. (2014), on the other hand, suggested that higher tibia Pb in women was inversely
associated with individual cognitive scores representing executive function and memory domains, and
was unexpectedly positively associated with immediate verbal memory domain. Associations with other
cognitive domains or the overall cognitive score were imprecise. Similarly, analyses from the NAS men
cohort with shorter versus extended follow-up periods indicated heterogeneous associations with global
and domain scores as well. Results from the analysis with shorter follow-up (Weisskopf et al., 2007)
suggested stronger inverse association between higher patella Pb and declines in the visuospatial and
visuomotor domains over time, but weaker and imprecise associations were observed for other domains.
In contrast, results from the extended analysis Farooqui et al. (2017) observed that higher patella Pb was
associated with faster longitudinal decline in MMSE (a measure of global cognition) and declines in the
language and memory domains, whereas associations with other cognitive scores or domains were
imprecise.
Recent studies of rodents with exposure resulting in mean BLLs <30 (ig/dL add to the evidence
informing the association of both short- and long-term Pb exposure during adulthood with measures of
learning and memory in rodents. While these studies are consistent with one another, toxicological
evidence for effects of Pb on cognitive function in adults remains limited., A few recent studies in
juvenile rodents also provide some support for the association between postnatal Pb exposure either
during adolescence or continuing into adulthood and cognitive impairment, specifically learning and
memory. Previous studies in nonhuman primates demonstrated that early life exposure to Pb may produce
cognitive impairment in adulthood. Hence, these findings add to the current evidence base suggesting
potential roles of both early and later life Pb exposures to produce cognitive function decrements in
adults. Additionally, animal, and in vitro studies lend biological plausibility to the association between
adult Pb exposure and adult cognitive impairment, showing that Pb has negative effects on neuronal
function and integrity, neurotransmission, and synaptic plasticity in regions of the brain associated with
learning and memory (Section 3.6).
Overall, the collective evidence is sufficient to conclude that there is a causal relationship
between Pb exposure and cognitive effects in adults. Recent prospective epidemiologic studies expand
and strengthen the previous body of evidence. These recent prospective studies include extended analyses
with longer follow-up periods, repeated measurements of cognitive outcomes, and adjustment for an array
of important potential confounders. Together, they provide compelling evidence for an association
between Pb exposure during various lifestages including childhood and decreased cognitive function in
adulthood. Discordant Pb associations across domains of cognitive function likely reflect differences in
the outcome pathophysiology rather than inconsistent results, and they do not detract from the strength of
the evidence overall. Recent evidence from animal studies supports the biological plausibility for the
effects of Pb exposure on cognitive function in adults and demonstrates that postnatal exposure to Pb
(either during adolescence or continuing into adulthood) may also negatively affect learning and memory.
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Table 3-9 Summary of evidence for a causal relationship between Pb exposure and cognitive effects in
adults
Rationale for Causality Determination3
Key Evidence13
References'3
Pb Biomarker
Levels Associated
with Effects0
Consistent findings from prospective epidemiologic
studies with relevant adult bone Pb levels or early
childhood BLLs suggesting significant association
of long-term and early childhood Pb exposure with
cognitive impairment and decline
Prospective analyses in NAS cohort of white men and NHS
cohort of white women found cognitive function decrements
over the 13 to 15 yr follow-up in association with patella or
tibia Pb levels.
Power et al. (2014)
Farooaui et al. (2017)
Mean patella Pb:
12.6 |jg/g
Mean tibia Pb: 10.5
pg/g
Mean patella Pb:
30.6 |jg/g
Mean tibia Pb:
21.6 |jg/g
Prospective analyses of childhood Pb exposure and long-term
cognitive impairments suggested persistent effects on
cognition during adulthood by showing lower IQ at young- and
mid-adulthood periods, and significant decline in IQ between
childhood and adulthood periods due to exposure to higher
childhood Pb levels.
Skerfvinq et al.
(2015)
Reuben et al. (2017)
Reuben et al. (2020)
Mean childhood (7-
12 yr) blood Pb: 3.4
pg/dL
Mean childhood (11
yr) blood Pb: 10.99
pg/dL
Models adjusted for various confounding factors including
baseline individual-level, socioeconomic, demographics,
behavioral, and clinical factors, as well as various
neighborhood level variables. The early childhood Pb
exposure studies also adjusted for parental education, HOME
scores, parent IQ and childhood IQ.
Supporting epidemiologic studies
Analysis of bone Pb and cognitive function among cases and
controls of PD found lower cognitive performance score with
increased tibia Pb among cases.
Weuve et al. (2013)
Mean tibia Pb: 4.4-
9.2 |jg/g (for age
groups)
Mean patella Pb:
5.9-15.2 |jg/g (for
age groups)
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Rationale for Causality Determination3
Key Evidence13
References'3
Pb Biomarker
Levels Associated
with Effects0
Cross-sectional analysis of concurrent blood Pb (along with
multi-metals) and cognitive function using NHANES data and
advanced modeling approach suggested significant inverse
association between concurrent BLL and cognitive outcomes.
Przvbvla et al. (2017)
Sasaki and
Carpenter (2022)
Geometric mean
blood Pb: 2.17 pg/dL
Mean blood Pb: 1.9
pg/dL
Models adjusted for socioeconomic and demographic factors,
education, health status, comorbidities, and co-exposure to
other metals.
Consistent evidence in animals with relevant
exposures
Recent evidence from animal studies supports the notion that
postnatal exposure to Pb (either during adolescence or
continuing into adulthood) negatively affects learning and
memory in rodents.
Mansouri et al.
(2012)
Mansouri et al.
(2013)
Sinqh et al. (2019)
Su et al. (2016)
Mean BLL: 8 pg/dL
Peak BLL: 11-19
pg/dL
Peak BLL: 28 pg/dL
Mean BLL: 8.4 pg/dL
Some uncertainty remains
Sex and age differences in bone kinetics and turnover may
contribute differences in biomarker Pb levels.
BLL = blood lead level; HOME = Health Outcomes and Measures of the Environment; IQ = intelligence quotient; NAS = Normative Aging Study; NHANES = National Health and
Nutrition Examination Survey; NHS = Nurses' Health Study; Pb = lead; PD = Parkinson's disease, yr = year.
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 20151.
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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3.6.2
Psychopathological Effects in Adults
The evidence assessed in the 2013 Pb ISA was sufficient to conclude that "a causal relationship is
likely to exist" between Pb exposure and psychopathological effects in adults. Cross-sectional studies in a
small number of distinct U.S. populations demonstrated associations of higher concurrent blood or tibia
Pb levels with self-reported symptoms of depression and anxiety in adults (Bouchard et al.. 2009; Raj an
et al., 2007; U.S. EPA, 2006). The examination of multiple exposures and outcomes in the available
studies does not provide a strong indication of biased reporting of psychopathological effects specifically
by adults with higher Pb exposures. In adults, Pb-associated increases in depression and anxiety were
found with adjustments for age, SES, and in the NAS, daily alcohol intake. The biological plausibility for
epidemiologic evidence was provided by observations of depression-like behavior in animals with dietary
lactational Pb exposure, with some evidence at relevant BLLs. In addition, Pb-induced changes in the
HPA axis and dopaminergic and GABAergic systems were demonstrated in animals. Overall, the
strongest evidence was from epidemiologic studies of adults without occupational Pb exposure, with
additional support from a few experimental animal studies; however, uncertainties related to residual
confounding of bone Pb associations by age in epidemiologic studies remained. Overall, recent studies
add to the evidence and generally support the findings from the 2013 Pb ISA.
3.6.2.1 Epidemiologic Studies of Psychopathological Effects in Adults
A limited number of epidemiologic studies evaluated in the 2006 Pb AQCD (U.S. EPA, 2006)
and 2013 Pb ISA (U.S. EPA, 2013) examined the relationship between blood or bone Pb levels and
psychopathological effects in adults. All of these studies were cross-sectional and most examined
occupationally exposed populations. In addition to the occupational studies, which provided consistent
evidence of positive associations between BLLs (mean levels >15 (ig/dL) and the prevalence of self-
reported symptoms of depression, anxiety, and tension, a prospective analysis of the NAS cohort (Raj an
et al., 2007) and a cross-sectional analysis of NHANES participants (Bouchard et al., 2009) reported
positive associations between much lower concentrations of concurrent blood (NHANES; mean ~ 6
(ig/dL) or bone (NAS) Pb levels and symptoms of depression and anxiety or prevalent major depressive
disorder, respectively. The findings from the NAS cohort are notable because tibia and patella Pb reflect
cumulative exposures that occur over several years to decades and thus serve as a retrospective
assessment of Pb exposure despite the cross-sectional study design.
Recent evidence includes several prospective cohort studies and cross-sectional analyses of Pb
exposure and general psychopathological effects or internalizing symptoms, as well as a case-control
study of schizophrenia. In general, recent prospective studies provide evidence of an association between
blood or bone Pb levels and psychopathological effects in adults. Results from cross-sectional studies are
inconsistent. Additionally, with blood or bone Pb levels, it is difficult to characterize the specific timing,
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duration, frequency, and level of Pb exposure that contributed to associations observed with cognitive
function. This uncertainty may apply particularly to assessments of BLLs, which in nonoccupationally
exposed adults reflect both current exposures and cumulative Pb stores in bone that are mobilized during
bone remodeling. Measures of central tendency for blood and bone Pb levels used in each study, along
with other study-specific details, including study population characteristics and select effect estimates, are
highlighted in Table 3-15E. An overview of the recent evidence is provided below.
A few recent prospective cohort studies report generally consistent evidence of associations
between blood or bone Pb levels and psychopathological effects in adults, although the results for specific
endpoints are not entirely consistent. In a recent prospective cohort study examining a subset of the NAS
cohort, Peters et al. (2011) used structural equation models to examine the interrelations of childhood and
adult SES, bone Pb levels, pessimism, and depression in older adults. After controlling for childhood and
adult SES through a combination of latent variables—including parental and participant education,
occupation, and home ownership—10 (ig/g higher tibia Pb levels measured prior to psychological
measurement were independently associated with a 0.3-unit (95% CI: 0.0, 0.6) higher score for pessimism
level on the Life Orientation Test. An independent association between bone Pb levels and depression
was not observed when controlling for pessimism (quantitative results not reported), but pessimism was
strongly associated with increased odds of depression (OR = 1.04 [95% CI: 1.02, 1.05] per 1 unit higher
pessimism level), indicating a potential mediating effect of pessimism on the relationship between bone
Pb levels and depression. Other recent prospective cohort studies examined the relationship between
childhood BLLs and a wider range of psychopathological effects (Reuben et al.. 2019; McFarlane et al..
2013). In an analysis of the Dunedin cohort in New Zealand, Reuben et al. (2019) reported that each 1-
(ig/dL higher BLL at age 11 was associated with 0.27-point (95% CI: 0.02, 0.51) higher standardized
general psychopathology scores (mean [SD]: 100 [15]) in early adulthood. In more specific analyses of
psychopathology components, the association with general psychopathology appeared to be driven by
positive associations with symptoms of internalizing behavior and thought disorder. These results are
somewhat consistent with a follow-up analysis of the Port Pirie cohort, a birth cohort from a South
Australian Pb-smelting town (McFarlane et al.. 2013). In this study, BLLs averaged over the first 7 years
of life were not associated with depressive symptoms measured during follow-up at ages 25 through 29.
However, the authors did report positive associations between BLLs and some internalizing behaviors in
women (e.g., social phobia, specific phobia, and anxiety problems). There were generally null
associations for the same outcomes in men. A notable limitation of this study is that attrition in the
original cohort led to a small sample size that was even further reduced in sex-stratified models, resulting
in limited power to detect an association. Additionally, due to high community exposure to Pb, the
participants in the Reuben et al. (2019) and McFarlane et al. (2013) studies had a high mean childhood
BLL (11.08 and 17.2 (ig/dL, respectively). The mean BLLs in these studies are not directly comparable to
other studies in this section, which used concurrent BLLs in adult populations that likely had higher past
exposures.
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Associations between BLLs and mood states were inconsistently observed in cross-sectional
studies of pregnant women in China (Li et al.. 2017) and Japan (Ishitsuka et al.. 2020). In a small study of
pregnant women in Shanghai, Li et al. (2017) observed nonlinear associations between BLLs and
depression, anxiety, and psychological distress in nonparametric models. Based on visual inspection of
the spline curves, the authors ran separate piecewise linear regression models for each outcome with a
knot at 2.57 (ig/dL. In the piecewise models, BLLs were associated with higher prevalence of depression,
anxiety, and psychological stress at levels below 2.57 (ig/dL, whereas smaller inverse associations were
observed above the knot. In contrast, a much larger cross-sectional study of pregnant women across Japan
noted that BLLs measured during middle or late pregnancy were not associated with increased odds of
Kessler Psychological Distress Scale (K6) scores greater than or equal to 5 or 13 (Ishitsuka et al.. 2020).
The authors used different cut points to account for potential differences in the optimal sensitivity-
specificity tradeoff for assessing depression in their study population. The contrasting results in these
studies are not readily explained by variations in BLLs, as Li et al. (2017) observed positive associations
at low BLLs in linear spline models, and the population analyzed by Ishitsuka et al. (2020) had a
geometric mean BLL <1 (ig/dL.
Other recent cross-sectional studies assessed the relationship between blood or bone Pb levels and
mood states measured by validated questionnaires or self-reported physician's diagnosis in a range of
study populations, including a population-based analyses of NHANES (Berk et al.. 2014) and KNHANES
participants (Nguyen et al.. 2022). older women participating in subcohorts of the NHS (Eum et al..
2012). and older adults from two communities selected using cluster-based sampling of communities in
Luan, China (Fan et al.. 2020). Similar to previously discussed studies, results from these analyses were
inconsistent. Fan et al. (2020) reported monotonic increases in odds of depression associated with
increasing blood Pb exposure quartiles. In this study, older adults in the highest quartile of exposure
(BLLs <3.06 (ig/dL) had just over twice the odds of depression as those with BLLs <2.03 (ig/dL (OR =
2.03 [95% CI: 1.23, 3.35]). In contrast, Eum et al. (2012). Berk et al. (2014). and Nguyen et al. (2022)
reported null associations between bone or blood Pb levels and anxiety or depression. However, in a
subgroup analysis restricting the study population to pre- and postmenopausal women taking hormone
replacement therapy (HRT), Eum et al. (2012) reported that increasing tibia Pb tertiles were associated
with monotonically increasing odds of phobic anxiety and lower scores on the Mental Health Index 5-
item (MHI-5; indicating worse depressive symptoms). The authors restricted the analysis by HRT status
to account for potential exposure measurement error in the non-HRT population, resulting from higher
variability in bone turnover. Notably, patella Pb was not associated with depressive symptoms in this
population and was inversely associated with phobic anxiety. Since tibia Pb has a longer half-life than
patella Pb, the results could be indicative of a long-term exposure window contributing to changes in
anxiety and depression. However, the subgroup analyses also had small sample sizes and thus a lack of
precision and higher probability of chance findings.
While most recent studies of psychopathological effects in adults examined internalizing
behaviors or general psychopathological effects, a small case-control study in China evaluated serum
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heavy metal levels in association with the risk of schizophrenia (Ma et al., 2019). The authors reported
that higher serum Pb levels were associated with higher odds of schizophrenia, but the association was
imprecise (OR = 3.15 [95% CI: 1.24, 7.99] per ng/mL higher BLL). Although the cases and controls were
matched on age and sex, it is unclear what, if any, other potential confounders were included in the
adjusted models. This finding was ostensibly consistent with a pooled analysis of two small cohorts
evaluated in the 2013 Pb ISA that observed an association between higher S-ALAD levels and increased
odds of schizophrenia spectrum disorder in adolescents and adults (Opler et al., 2008). However, this
pooled analysis had a number of limitations that precluded any conclusions regarding a relationship
between Pb exposure and schizophrenia, including the lack of direct measurements of Pb biomarker
levels and limited consideration for potential confounding. As noted previously, Reuben et al. (2019)
reported a positive association between childhood BLLs and symptoms of thought disorder in early
adulthood. However, the metric for thought disorders included factor loadings for obsessive compulsive
disorder and mania in addition to schizophrenia, making it difficult to distinguish an independent
relationship between BLLs and schizophrenia.
3.6.2.1.1 Summary
A limited number of cross-sectional studies evaluated in the 2013 Pb ISA (U.S. EPA, 2013)
provided consistent evidence of positive associations between blood and bone Pb levels and the
prevalence of self-reported symptoms of depression, anxiety, and tension. Recent prospective analyses
provide additional support for a positive association between bone and BLLs and psychopathological
effects in older adults, although results from cross-sectional studies are inconsistent.
3.6.2.2 Toxicological Studies of Psychopathological Effects in Adults
No studies in the 2013 Pb ISA evaluated the effect of adult-only exposure on anxiety and
depression-like behaviors in animals. Nevertheless, developmental studies consistently supported an
effect of Pb exposure on these endpoints in adult animals (U.S. EPA, 2013). Of particular importance, the
increased reactivity to errors and reward omission reported by Beaudin et al. (2007) and Stangle et al.
(2007) extended to adulthood, well after postnatal exposure was terminated. Furthermore, Pb exposure
during adolescence reduced immobility on the FST in adult rats (Stewart et al., 1996).
A few recent studies have investigated adult-only exposure and anxiety-like outcomes using the
OFT and EPM. Similar to developmental exposures, adult male mice displayed anxiety-like behavior (i.e.,
increased time spent in closed arms) in the EPM following 6 weeks of Pb exposure given via oral gavage
(mean BLLs 7.1 (ig/dL) (Al-Qahtani et al., 2022). Singh et al. (2019) also found that oral gavage of Pb for
90 days decreased the amount of time rodents spent in the open arms of the EPM compared with control
animals. Another study of long-term adult exposure (126 days) to Pb in Wistar rats found increased
rearing and grooming, but not sniffing, in males, and no significant effects in females in the OFT
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(Mansouri et al.. 2013). The same research group found that Pb did not affect these endpoints (i.e.,
rearing, sniffing) following shorter-term exposure (30 days) in adult animals (Mansouri et al.. 2012).
Studies that employed developmental exposure paradigms are discussed in more detail in Section
3.5.4.2. However, they provide consistent evidence to support the notion that the effects of developmental
exposures may persist into adulthood. In one study of gestational exposure, there were significant
decreases in exploratory behaviors in the OFT and hole board test in Wistar rats at 4 months old (BLLs
peaked at 12 (ig/dL and decreased to 6 (ig/dL by 4 months) (Basha and Reddv. 2015). A similar study
utilizing postnatally exposed rats found that some measures of decreased exploratory behavior in the OFT
and hole board test persisted until 18 months, when the study was terminated (Basha et al.. 2014). The
mean BLL at PND 45 in this study was high (50 (ig/dL) but had decreased to 11 (ig/dL by 18 months.
Several additional studies investigated lifetime Pb exposure (gestation through 6-12 months), and all
found significant treatment effects on EPM or FST behavior during adulthood (Shvachiv et al.. 2020.
2018; Abazvan et al.. 2014; Corv-Slechta et al.. 2013).
Toxicological studies also provide biological plausibility to support a connection between
exposure to Pb and schizophrenia. As discussed in the 2013 Pb ISA, antagonists of NMDAR's glycine
site have been shown to exacerbate schizophrenia symptoms in affected individuals and induce a
schizophrenic phenotype in unaffected subjects (Covle and Tsai. 2004). Previous studies have shown that
Pb is a potent allosteric inhibitor at NMDARs (Hashcmzadch-Gargari and Guilarte. 1999; Guilarte. 1997).
A recent study found that developmental Pb exposure (BLLs of 22 (ig/dL at PND 50) reduced the number
of parvalbumin-positive GABAergic interneurons in the median prefrontal cortex and hippocampus and
induced subcortical dopaminergic hyperactivity, consistent with studies of schizophrenic patients
(Stansfield et al.. 2015; Volman et al.. 2011). Pb exposure may also affect DISC 1, a gene-protein pair
associated with increased susceptibility to schizophrenia and other mental disorders. You et al. (2012)
found that expression of the DISCI protein was increased in the hippocampus of rats exposed to Pb
during gestation and lactation. Abazvan et al. (2014) utilized a transgenic mouse model expressing mutant
DISCI (mDISCl) to evaluate a potential gene-environment interaction using lifetime Pb exposure. Pb-
exposed mDISC 1 mice exhibited behavioral and structural abnormalities consistent with schizophrenia,
which were not found in unexposed mDISC 1 mice or Pb exposure regular mice (i.e., heterozygous for
mDISCl but phenotypically normal).
3.6.2.2.1 Summary
Toxicologic studies providing support for adult psychopathological effects in the previous ISA
used developmental exposure paradigms. Recent developmental exposure studies were consistent with the
previous evidence and were predominantly focused on anxiety-like behaviors. Multiple studies
demonstrated the persistence of these effects into adulthood (up to 1.5 years), in some cases long after
termination of Pb exposure. A few recent studies focused on adult-only exposures found some
associations with anxiety-like behavior after 42-126 days of exposure but not following a 30-day
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exposure; additional studies are needed to strengthen this line of evidence. Two recent studies also
provided further biological plausibility support for an NMDAR-mediated association between Pb
exposure and schizophrenia.
3.6.2.3 Relevant Issues for Interpreting the Evidence Base
3.6.2.3.1 Concentration-Response Function
One recent study used a nonparametric model to assess the C-R relationship between BLLs and
depressive symptoms in pregnant women (Li et al.. 2017). The authors observed nonlinear associations
between BLLs and depression, anxiety, and psychological distress scores and used the nonparametric
models to determine knots for a piecewise linear regression model. The subsequent models indicated
positive associations between BLLs and depression, anxiety, and psychological stress at levels below 2.57
(ig/dL, whereas smaller inverse associations were observed for BLLs above 2.57 (ig/dL. Other cross-
sectional studies reported inconsistent evidence of associations despite evaluating populations with low
mean BLLs. However, in studies analyzing BLLs in adult populations with higher past exposures, it is a
challenge to ascertain the level, timing, frequency, and duration of Pb exposure that contributed to
observed associations.
3.6.2.3.2 Potentially At-Risk Populations
A few of the recent epidemiologic studies detailed in this section evaluated populations that are
potentially at-risk for Pb-related health effects. The conclusions that can be drawn from these analyses are
limited. A prospective analysis of young adults reported sex-specific associations between childhood
BLLs and internalizing symptoms in early adulthood (McFarlanc et al.. 2013). The observed associations,
which were only present in stratified models including women, were extremely imprecise due to a small
sample size that was even further reduced by stratification. The small sample size in this study reduced
the statistical power to detect an association and the likelihood that an observed result reflects a true
effect, making it difficult to draw firm conclusions on these sex-specific comparisons. Additionally, two
cross-sectional studies examining depressive symptoms in pregnant women observed inconsistent
evidence of an association with BLLs (Ishitsuka et al.. 2020; Li et al.. 2017).
3.6.2.3.2.1 Lifestages
Toxicological studies provide consistent evidence that developmental and lifetime exposures to
Pb can lead to increases in anxiety-like behaviors. Comparatively fewer studies have investigated the
effects of adult-only exposure, but some effects have been demonstrated. Epidemiologic studies provide
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some supporting evidence for the importance of developmental and cumulative exposures. In a recent
prospective cohort study, Peters et al. (2011) reported that increased depressive symptoms in older adults
were associated with tibia Pb levels, a measure of cumulative exposure. Other recent prospective analyses
observed associations between childhood BLLs and increased internalizing symptoms in young adults
(Reuben et al.. 2019; McFarlane et al.. 2013). Additionally, (Reuben et al.. 2019) also observed positive
associations between childhood BLLs and internalizing symptoms at the time of BLL testing, which is
coherent with toxicological evidence that suggests the persistence of developmental effects into
adulthood. Given the uncertainties regarding potentially higher historical exposures in adults, the cross-
sectional epidemiologic studies evaluated in this section are less suited to address the importance of
concurrent exposures.
3.6.2.3.3 Confounding
The studies evaluated in this section controlled for a range of potential confounding variables that
may be associated with both Pb exposure and psychopathological effects, including age, sex, SES factors,
and marital status (see Table 3-15E).
3.6.2.4 Summary and Causality Determination: Psychopathological Effects in Adults
The 2013 Pb ISA (U.S. EPA, 2013) concluded that the available evidence was sufficient to
conclude that "a causal relationship is likely to exist" between Pb exposure and psychopathological
effects in adults. This causality determination was based on a small body of epidemiologic evidence that
demonstrated consistent positive associations between concurrent blood or bone Pb levels and self-
reported symptoms of depression, anxiety, and panic disorder in large studies of adults (i.e., NHANES,
NAS). The epidemiologic evidence was supported by coherence in animal toxicological studies that
demonstrated depression-like behavior and emotionality in rodents exposed to dietary lactational Pb with
or without additional post-lactational exposure. Epidemiologic associations were observed in study
populations of young (20-39 years old) and older (44-98 years old) adults. Because of the cross-sectional
design of the epidemiologic studies, there was uncertainty regarding the temporal sequence between Pb
exposure and psychopathological symptoms in adults. This uncertainty is somewhat reduced with results
for tibia Pb since it is an indicator of cumulative Pb exposure. Nonetheless, because these studies included
adults with likely higher past Pb exposures, uncertainties exist regarding the Pb exposure level, timing,
frequency, and duration contributing to the associations observed with blood or bone Pb levels. An
uncertainty in the toxicological evidence base was the limited number of studies that administered
exposures resulting in BLLs that are not relevant to humans. Recent epidemiologic and toxicological
evidence continues to link Pb exposure to psychopathological effects in adults, though some uncertainties
still remain. The key evidence, as it relates to the causality determination, is presented in Table 3-10 and
Table 3-11 and summarized below.
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Recent evidence from prospective epidemiologic studies provides further support for positive
associations between Pb exposures and pathological effects, including increased internalizing symptoms.
Specifically, a study of older men in the NAS cohort provides evidence of an (indirect) association
between Pb exposure and depression in adults (Peters et al., 2011), and an analysis of a cohort in New
Zealand similarly reported that increased childhood BLLs were associated with increased internalizing
symptoms in young adults (Reuben et al., 2019). Together, these studies address an uncertainty from the
previous ISA regarding the temporality of the exposure and outcome. Another recent prospective study of
young adults from the Port Pirie cohort reported null associations between childhood BLLs and
depression in adults but noted several sex-specific associations between BLLs and other internalizing
symptoms in young women (McFarlane et al., 2013). The small analytic sample of the stratified models
used in this analysis reduces the likelihood of detecting a true effect. Notably, supporting evidence from
recent cross-sectional epidemiologic studies conducted in diverse populations is largely inconsistent.
However, these studies are less informative given the limitations of the study design.
The epidemiologic evidence is supported by coherence with results from an expanded number of
toxicological studies conducted at BLLs relevant to humans. Recent toxicological studies examine
multiple exposure windows and provide strong support for Pb-induced anxiety-like behaviors following
developmental and cumulative exposures. Multiple studies demonstrate the persistence of these effects
into adulthood (up to 1.5 years), in some cases long after termination of Pb exposure. The evidence for
effects resulting from adult-only exposures is more limited, though there is some evidence for an increase
in anxiety-like behavior following 42-126 days of exposure but not following a 30-day exposure. The
2013 Pb ISA also highlighted Pb-induced changes in the dopaminergic and GABAergic systems and the
HPA axis, which underlie biological plausibility for the changes in mood and emotional state that have
been observed in epidemiologic and toxicological studies (U.S. EPA, 2013). Recent studies continue to
demonstrate changes in corticosterone and glucocorticoid receptors (i.e., HPA axis changes (Cory-Slechta
et al., 2012)) and the dopaminergic system (Section 3.4.2.2).
In addition to studies of depression, anxiety, and mood-related disorders, some recent studies
examined the relationship between Pb exposure and schizophrenia. A recent case-control study in China
reported a positive, but imprecise association between serum Pb levels and schizophrenia prevalence in
adults (Ma et al„ 2019). Because serum Pb was measured after schizophrenia was diagnosed, the results
do not establish temporality between exposure and outcome. Additionally, the reported analytic
methodology does clarify the confounding variables considered outside of the age- and sex-based
matching of cases and controls. Recent toxicological studies provide biological plausibility to support a
connection between exposure to Pb and schizophrenia. Consistent with evidence from the 2013 Pb ISA,
two recent studies support Pb-induced pathophysiological features in rodents consistent with
schizophrenia, likely through inhibition of NMDAR activity. Although the toxicological evidence
presents a biologically plausible pathway through which exposure to Pb could lead to schizophrenia, the
limited quantity and quality of the epidemiologic evidence precludes meaningful consideration of this
endpoint in the causality determination for Pb exposure and psychopathological effects.
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Overall, the collective evidence is sufficient to conclude that there is likely to be a causal
relationship between Pb exposure and psychopathological effects in adults. The strongest evidence
comes from a limited number of recent prospective epidemiologic studies that add to previous evidence of
a positive association between bone or BLLs and psychopathological effects in older adults and addresses
prior uncertainties regarding the temporality of exposure and outcome. Recent toxicological studies
strengthen the overall evidence base, providing further support for anxiety-like behaviors following
developmental and cumulative exposures that result in BLLs that are relevant to humans. Despite
generally consistent evidence from prospective epidemiologic studies that Pb is associated with general
internalizing behavior scores or some components of internalizing behavior, the evidence from these
limited number of studies is not consistent for any single component. Although these inconsistencies may
reflect differences in outcome pathophysiology rather than inconsistent results, there is remaining
uncertainty given the limited body of evidence. The key evidence, as it relates to the causal framework, is
summarized in Table 3-10.
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Table 3-10 Summary of evidence for a likely to be causal relationship between Pb exposure and
psychopathological effects in adults
Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels
Associated with
Effects0
Limited epidemiologic evidence from Prospective analyses reported positive associations between tibia Pb levels
high-quality prospective cohort studies and depressive symptoms in older men, mediated by pessimism; and
with relevant bone Pb levels childhood BLLs and psychopathology scores in early adulthood.
Some supporting evidence from a prospective analysis reported imprecise
positive associations between childhood BLLs and prevalence of social
phobia, specific phobia, PTSD, anxiety problems, somatic problems, and
antisocial personality problems in young adult women.
Peters et al. Mean: 20.6 |jg/g
(2011)
Reuben et al.
(2019)
Mean: 11.08 pg/dL
McFarlane et Mean: 17.2 |jg/dL
al. (2013)
Consistent evidence in animals with
relevant exposures
Increased anxiety-like behaviors in adulthood (up to 18 mo) following
developmental or lifetime Pb exposure provides coherence with epidemiologic
evidence.
Section 3.5.4.2
Peak blood Pb after
lifetime exposure: 7-24
pg/dL
Peak blood Pb after
developmental exposure:
12-50 |jg/dL
Inconsistent supporting evidence from
cross-sectional epidemiologic studies
with relevant blood and bone Pb levels
A limited body of cross-sectional studies provides inconsistent evidence of
associations between generally lower blood and bone Pb levels and
depression and anxiety.
Section 3.6.2.1 Mean/median range
across studies:
Bone: 10.3-12.5 |jg/dL
Blood: 0.58-3.97 pg/dL
Uncertainty regarding potential
confounding
Most studies included adjustment for age, sex, SES factors, and marital status,
Table 3-15E but did not consider use of antidepressants
BLL = blood lead level; mo = months; Pb = lead; PTSD = post-traumatic stress disorder; SES = socioeconomic status.
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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3.6.3
Sensory Organ Function in Adults
The 2013 Pb ISA included separate causality conclusions for auditory and visual function. This
ISA combines these categories and makes one causality determination for Sensory Organ Function (see
Section 3.6.3.5). Recent studies are summarized in Table 3-16E and Table 3-16T. An overview of the
recent evidence is provided below.
3.6.3.1 Auditory Function
The evidence assessed in the 2013 Pb ISA was "suggestive of a causal relationship" between Pb
exposure and auditory function decrements in adults (U.S. EPA, 2013). The strongest evidence was
provided by the analysis of NAS men, which revealed associations between higher tibia Pb levels and a
higher rate of elevations in hearing threshold over 20 years (Park et al., 2010). Findings demonstrating
decreased auditory evoked potentials in animals provided biological plausibility for the observations in
this epidemiologic study, but uncertainties related to effects on auditory function in adult animals with
relevant Pb exposures remained.
3.6.3.1.1 Epidemiologic Studies of Auditory Function
Several recent epidemiologic studies examined the association between Pb exposure and
decrements in auditory function in adults (Tu et al., 2021; Yin et al., 2021; Wang et al„ 2020; Kang et al.,
2018; Choi and Park, 2017; Shiue, 2013; Choi et al., 2012). The findings generally supported an
association between Pb exposure and hearing loss in adults. These studies are described below.
Most studies were cross-sectional, using data from NHANES (Tu et al„ 2021; Shiue, 2013; Choi
et al., 2012) and KNHANES (Kang et al„ 2018; Choi and Park, 2017). Choi et al. (2012) and Tu et al.
(2021) measured blood Pb in adult NHANES participants (20-69 years) in 1999-2004 and 2011-2012,
respectively. Choi et al. (2012) observed an increased likelihood of hearing loss per doubling of blood Pb
(OR = 1.09 [95% CI: 0.95, 1.26]) as well as an increased percent change in hearing threshold (% change
= 5.41 [95% CI: 2.12, 8.81]). Similar findings were noted in quintile analyses, with higher blood Pb
quintiles having a greater magnitude of effect when compared with the lowest quintile (0.20-0.80 (ig/dL)
(Choi et al., 2012). Tu et al. (2021) measured hearing loss at speech frequency and at high frequency. In
quartile analyses, the magnitude of effect increased with each blood Pb quartile for each type of hearing
loss (Tu et al., 2021). Compared with the lowest quartile (<0.07 |ig/dL). the highest blood Pb quartile
(>0.16 (ig/dL) was positively associated with speech-frequency hearing loss (OR = 1.46 [95% CI: 0.81,
2.64]) and high-frequency hearing loss (OR = 1.98 [95% CI: 1.27, 3.10]). In sex-stratified analyses, the
direction of effect remained but the magnitude of effect was greater among males. In age-stratified
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analyses, a positive association remained for ages <35 years and 35-52 years for both types of hearing
loss; however, ORs were generally imprecise for the younger age group (<35 years). On the contrary,
there appeared to be an inverse association between BLLs and hearing loss in the oldest age group (>52
years) (Tu et al.. 2021). Also using NHANES data, Shiue (2013) measured Pb exposure in urine and did
not find an adverse association with self-reported hearing among older adults (>50 years). For self-
reported ear ringing, urinary Pb had a slightly positive association (Shiue. 2013). In KNHANES, Choi
and Park (2017) measured speech- and high-frequency hearing loss in adolescents (12-19 years) and
adults (20-87 years). Hearing loss was defined as pure-tone average >25 dB in adults. For each doubling
of blood Pb, there was an increased likelihood of speech-frequency hearing loss (OR = 1.15 [95% CI:
0.94, 1.41]) and high-frequency hearing loss (OR = 1.30 [95% CI: 1.08, 1.57]). In quartile analyses, the
magnitude of effect increased with each blood Pb quartile when compared with the lowest quartile (Choi
and Park. 2017). In another analysis conducted in the KNHANES population, Kang et al. (2018) observed
an association between BLLs and hearing impairment in adults (20-87 years). In quartile analyses, the
magnitude of effect for high-level frequency hearing impairment was greatest in the highest blood Pb
quartile compared with the lowest blood Pb quartile in males (OR = 1.63 [95% CI: 1.16, 2.29]) as well as
in females (OR = 1.50 [1.03-2.20]). For low-frequency hearing impairment, the associations were less
consistent by quartile and more attenuated (Kang et al.. 2018).
In a meta-analysis of studies from Iran, Korea, China, and the United States, Yin et al. (2021)
observed consistent positive associations between Pb exposure and any hearing loss (combined OR per
unit increase in Pb = 1.42 [95% CI: 1.22, 1.67]), low-frequency hearing loss (combined OR = 1.31 [95%
CI: 1.17, 1.47]), and high-frequency hearing loss (combined OR= 1.96 [95% CI: 1.48, 2.60]). When
stratified by age group, the association persisted in adults (>20 years; combined OR per unit increase in
Pb = 1.34 [95% CI: 1.18, 1.52]) (Yin et al.. 2021). Despite these results, there are still some
inconsistencies in the recent literature. In a case-control study of adults in China who participated in a
survey of hearing loss, Wang et al. (2020) did not observe an association with blood Pb before and after
adjusting for workplace noise exposure.
Summary
The strongest evidence described in the 2013 Pb ISA was provided by the analysis of NAS men
for associations of higher tibia Pb level with a higher rate of elevations in hearing threshold over 20 years
(Park et al.. 2010). Several recent cross-sectional analyses of NHANES and KNHANES generally support
an association of Pb exposure (i.e., concurrent BLLs) with hearing loss; however, recent studies are not
entirely consistent.
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3.6.3.1.2 Toxicological Studies of Auditory Function
Recent animal studies on the auditory effects of Pb exposures have investigated exposures
beginning during development (postnatal or adolescent), discussed in Section 3.5.6.1.2. In particular,
Carlson et al. (2018) did not detect any significant changes in BAEP in 4-month-old adult mice with very
low BLLs (3 (ig/dL). The strongest evidence for adult effects of Pb exposure was presented in the
previous ISA (U.S. EPA, 2013). Lifetime Pb exposure was found to increase hearing thresholds and
latencies in BAEP in adult monkeys (aged 8-13 years) (Rice, 1997; Lilienthal and Winneke, 1996).
Moreover, Laugh 1 in et al. (2009) detected small nonsignificant shifts in auditory threshold in 13-year-old
Rhesus monkeys following gestational or postnatal Pb exposure. However, these effects were
demonstrated at higher BLLs than are relevant to this ISA (33-150 (.ig/dL).
3.6.3.2 Visual Function
The evidence pertaining to visual function assessed in the 2013 Pb ISA was limited. A case-
control study found higher Pb in retinal tissue from macular degeneration cases but lacked rigorous
statistical analysis and examination of potential confounding. Studies in adult animals showed differential
effects on ERGs, depending on the timing and concentration of exposure. Because the available
epidemiologic and toxicological evidence was of insufficient quantity, quality, and consistency, the 2013
Pb ISA concluded the "evidence is inadequate to determine that a causal relationship exists between Pb
exposure and visual function decrements in adults."
3.6.3.2.1 Epidemiologic Studies of Visual Function
Only a few epidemiologic studies examined the association between Pb exposure and decrements
in visual function in adults (Paulsen et al.. 2018; Fillion et al.. 2013; Shiue. 2013). Fillion et al. (2013)
measured contrast sensitivity (cpd) and acquired color vision loss (CCI) in adolescents and adults (15-66
years) in Brazil. Blood Pb exposure was negatively associated with the intermediate spatial frequency of
contrast sensitivity (12 cycles/degree); however, results varied by spatial frequency. For CCI, there was a
slightly positive association with blood Pb, but the effect estimate was imprecise (Fillion et al.. 2013). In
another study of contrast sensitivity, Paulsen et al. (2018) used the Pelli-Robson letter sensitivity chart
and did not observe an association with blood Pb in a U.S.-based cohort (HR for blood Pb >2.06 |ig/L
versus. <2.06 |ig/L = 0.91 [95% CI: 0.69, 1.18]). Visual function has also been measured using self-
reported eyesight. Using NHANES data, Shiue (2013) measured Pb exposure in urine and did not find an
association with self-reported visual impairment among older adults (>50 years).
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Summary
The epidemiologic evidence pertaining to the association of Pb exposure with visual function in
adults remains limited. A small number of recent studies examining contrast sensitivity or acquired color
vision loss found inconsistent results for associations with blood or urine Pb level (Paulsen et al., 2018;
Fillion et al., 2013; Shiue. 2013).
3.6.3.2.2 Toxicological Studies of Visual Function
No recent PECOS-relevant studies have evaluated the effects of Pb exposure on visual function.
Section 3.5.6.2.2 summarizes the literature discussed in the 2013 Pb ISA (U.S. EPA, 2013).
3.6.3.3 Olfactory Function
The 2013 Pb ISA did not assess any evidence on the relationship between Pb exposure and
olfactory function in adults (U.S. EPA, 2013).
3.6.3.3.1 Epidemiologic Studies of Olfactory Function
In the Heinz Nixdorf Recall Study (HNRS) in Germany, male participants were recruited in
2000-2003 and followed up in 2011-2014 (Casjens et al., 2018). Casjens et al. (2018) examined the
effect of Pb exposure on odor identification using the Sniffin' sticks odor identification test of 12 odors.
Participants were classified as normosmic (identified >9 odors), hyposmic (identified 7-9 odors), and
functionally anosmic (identified <7 odors). Compared with the lowest BLL at baseline (<5 |ig/dL). the
highest BLL (>9 (ig/dL) was associated with impaired odor identification (proportional OR = 1.96 [95%
CI: 0.94, 4.11]). Similar results were observed using BLLs measured at follow-up (proportional OR =
1.57 [95% CI: 0.47, 5.19]).
3.6.3.4 Relevant Issues for Interpreting the Evidence Base
3.6.3.4.1 Potentially At-Risk Populations
Sex
Tu et al. (2021) measured hearing loss at speech frequency and at high frequency. In quartile
analyses, the magnitude of effect increased with each blood Pb quartile for each type of hearing loss (Tu
et al.. 2021). In sex-stratified analyses, the direction of effect remained but the magnitude of effect was
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greater among males. Kang et al. (2018) also observed an association between BLLs and hearing
impairment in adults. In quartile analyses, the magnitude of effect for high-level frequency hearing
impairment was greatest in the highest blood Pb quartile compared with the lowest blood Pb quartile, with
similar associations observed in males (OR= 1.63 [95% CI: 1.16, 2.29]) and females (OR= 1.50 [1.03-
2.20]).
3.6.3.5 Summary and Causality Determination: Sensory Organ Function in Adults
In the 2013 Pb ISA, causality determinations were separately determined for auditory and visual
function in adults, while no studies of olfactory function were evaluated (U.S. EPA, 2013). For auditory
function in adults, the evidence was suggestive of, but not sufficient to infer, a causal relationship, based
primarily on findings from a few epidemiologic studies. Similar to the conclusion for children, the
evidence relating to visual function in adults was inadequate to determine if a causal relationship exists.
In the current ISA, the evidence for auditory, visual, and olfactory function are evaluated together,
forming a single causality determination for sensory organ function.
The strongest evidence described in the 2013 Pb ISA was provided by the analysis of NAS men
for associations between higher tibia Pb levels and a higher rate of elevations in hearing threshold over 20
years (Park et al., 2010). Several recent cross-sectional analyses of NHANES and KNHANES generally
support an association of Pb exposure (i.e., concurrent BLLs) with hearing loss (Tu et al., 2021; Kang et
al„ 2018); Choi and Park (2017); (Choi et al„ 2012); however, recent studies are not entirely consistent
(Wang et al., 2020). Hearing loss and altered responses on BAEPs in adult nonhuman primates and
rodents following lifetime or developmental Pb exposure have been demonstrated at BLLs as low as 29
(ig/dL (Jamesdaniel et al., 2018; Laughlin et al., 2009; Rice, 1997). The few studies that investigated
BLLs from 3-8 (ig/dL did not report altered BAEP in rodents, though effects on auditory processing may
occur at these lower exposure levels (Liu et al., 2019; Carlson et al„ 2018; Zhu et al„ 2016).
The epidemiologic and experimental animal evidence pertaining to the association of Pb exposure
with visual function in adults remains limited. Toxicological studies have demonstrated biological
plausibility for Pb-induced effects on vision, including dysfunction of subcortical visual neurons, visual
processing areas, and retinal development (U.S. EPA, 2013). A small number of recent studies examining
contrast sensitivity or acquired color vision loss in humans found inconsistent results for associations with
blood or urine Pb level (Paulsen et al„ 2018; Fillion et al„ 2013; Shiue, 2013). Deficits in visual temporal
acuity have been demonstrated in adult nonhuman primates, although peak exposure levels are higher
than considered relevant for this assessment (Rice, 1998). Altered responses to ERGs have been detected
at a wide range of BLLs, but the direction of this effect (i.e., supernormal or subnormal responses) is
overall inconsistent (Fox et al„ 2008; Rothenberg et al„ 2002; Fox et al„ 1997).
Olfactory function was not discussed in the 2013 Pb ISA. Recently, baseline BLL was associated
with reduced odor identification in a prospective analysis of the German HNRS (Casjens et al., 2018).
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Overall, the evidence is suggestive of, but not sufficient to infer, a causal relationship between
Pb exposure and sensory function in adults. This determination is supported by generally consistent
prospective and cross-sectional analyses demonstrating Pb-associated hearing loss in adults. The few
experimental animal studies available provide coherence for this endpoint when BLLs are greater than 29
(.ig/dL. but not at BLLs more relevant to current human exposures. Human and experimental animal
studies investigating the effect of Pb on visual and olfactory function are limited and inconsistent.
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Table 3-11 Summary of the evidence that is suggestive of, but not sufficient to infer, a causal relationship
between sensory function in adults
Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with
Effects0
Auditory Function
Generally consistent
associations observed in
multiple epidemiologic
studies.
Prospective study found association
between tibia Pb level and higher rate of
increase in hearing threshold over 23 yr in
males enrolled in the NAS.
Park etal. (2010)
Tibia Pb mean: 22.5 |jg/g, measured near
end of follow-up
Cross-sectional analyses of NHANES and
KNHANES find Pb-associated effects on
hearing loss
Shiue (2013)
Choi et al. (2012) Tu et al. (2021)
Kana etal. (2018)
Choi and Park (2017)
Uncertainty at relevant
exposure levels in
experimental animal studies
Hearing loss and altered responses on
BAEPs in adult nonhuman primates and
rodents demonstrated
Rice (1997)
Lauqhlin et al. (2009)
Jamesdaniel et al. (2018).
Lifetime or developmental Pb exposure
>29 |jg/dL
No altered BAEP in rodents, though
effects on auditory processing may occur
Carlson et al. (2018)
Zhu etal. (2016)
Liu et al. (2019)
3-8 |jg/dL
Visual Function
Inconsistent results across
limited epidemiologic studies
Associations of blood or urine Pb level
with contrast sensitivity or acquired color
vision loss were inconsistent
Shiue (2013)
Paulsen et al. (2018)
Fillion et al. (2013)
Limited evidence from
experimental animal studies
Deficits in visual temporal acuity
demonstrated in adult nonhuman
primates
Rice (1998)
>30 |jg/dL
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Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with
Effects0
Biological plausibility
demonstrated
Dysfunction of subcortical visual neurons,
visual processing areas, and retinal
development
(U.S. EPA, 2013)
Olfactory Function
Single study indicates
association
Impaired odor identification associated
with BLL in a single study
Casiens et al. (2018)
>9 vs. <5 |jg/dL
BAEP = brainstem auditory evoked potentials; BLL = blood lead level; KNHANES = Korea National Health and Nutrition Examination Survey; NAS = Normative Aging Study;
NHANES = National Health and Nutrition Examination Survey; Pb = lead; yr = year(s).
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 20151.
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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3.6.4
Neurodegenerative Diseases
The 2013 Pb ISA concluded that the evidence was "inadequate to determine that a causal
relationship exists" between Pb exposure and neurodegenerative diseases (U.S. EPA, 2013). Evidence
was inconclusive for amyotrophic lateral sclerosis (ALS) (see Section 4.3.9.2 of (U.S. EPA, 2013)) and
Alzheimer's disease (AD; see Section 4.3.9.1 of (U.S. EPA. 2013)); however, a few case-control studies
each found higher BLLs in adults with essential tremor and higher bone Pb levels in adults with PD
(Weisskopf et al„ 2010; U.S. EPA, 2006).The evidence was considered inconclusive overall due to a
limited number of studies, the potential for reverse causation (specifically for case-control studies in
which the reduced physical activity among cases could result in greater bone turnover and greater release
of Pb from bones into blood as compared with controls), and limited consideration for potential
confounding factors.
Recent epidemiologic studies indicated potential relationships between Pb exposure and some
neurodegenerative disease endpoints among non-occupational cohorts. The strongest evidence in the
current review includes well-designed studies of ALS and PD outcomes. Findings for AD, tremor, and
motor function were inconclusive. Measures of central tendency for Pb biomarker levels used in each
study, along with other study-specific details, including study population characteristics and select effect
estimates, are highlighted in Section 3.7, Table 3-17E. A large number of toxicological studies adds to the
evidence suggesting that developmental exposure to Pb increases the expression of pathophysiological
markers of AD, including amyloid beta (A|3) peptides, tau, and phosphorylated tau (p-tau), at lower BLLs
than were investigated in the previous ISA (<10 (.ig/dL). Toxicological studies investigating potential
associations between Pb exposure and PD, ALS, and essential tremor remain limited in this review.
However, toxicological evidence for PD and ALS are supported by some studies in the 2013 review that
provided pathophysiological evidence for Pb-induced decreases in dopaminergic cell activity in the
substantia nigra, which can contribute to PD development, and Pb exposure affecting neurophysiologic
changes associated with ALS.
3.6.4.1 Epidemiologic Studies of Neurodegenerative Diseases
3.6.4.1.1 Alzheimer's Disease
MMSE is a widely used screening tool for AD and other types of dementia. Lower scores on
MMSE were consistently associated with higher bone Pb levels, which indicated long-term or cumulative
exposure to Pb, in the NAS studies assessed in the 2013 Pb ISA (Wang et al.. 2007; Weisskopf et al..
2004; Wright et al.. 2003). There was heterogeneity in the results of studies that examined associations
between MMSE scores and BLLs in adults (Weuve et al.. 2006; Nordberg et al.. 2000). Blood Pb is
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generally considered a marker of recent exposure; however, in studies of adults, blood Pb level also
reflects Pb that is mobilized from the bone introducing uncertainty and complicating the interpretation of
cross-sectional studies that assess exposure using concurrent blood Pb level. Evidence regarding the
association of Pb exposure with clinical diagnosis of AD was limited to studies which did not find
associations with higher occupational exposure to Pb (Graves et al.. 1991) or higher Pb concentration in
the brains (Haraguchi et al.. 2001) in AD cases compared with unaffected controls. The latter studies were
limited because of their case-control designs, which may be subject to reverse causation where AD leads
to higher Pb levels, limited consideration for potential confounding, and because of the potential
misalignment with relevant exposure window.
Recent studies add to the evidence base, including an analysis of the NAS cohort that examined
the association of bone Pb biomarkers with cognitive impairment, including MMSE score, Farooqui et al.
(2017). and two studies that examined the association of blood Pb biomarkers with the clinical endpoints
of AD risk or AD mortality (Horton et al.. 2019; Yang et al.. 2018) in non-occupational cohorts (Table
3-16T). Among the older male participants in the NAS, higher patella Pb concentration (IQR: 21 |ig/g)
was associated with increased risk (HR: 1.10, 95% CI: 0.99, 1.21) ofhaving an MMSE score below 25
(threshold that represent cognitively not normal or at risk for dementia), while less support was observed
for an association with tibia Pb concentration (HR: 1.03, 95% CI: 0.88, 1.22) (Farooqui et al.. 2017).
Studies that specifically assessed clinically diagnosed AD or mortality did not provide strong evidence of
an association. A case-control study by Yang et al. (2018) included participants from clinical settings in
Taiwan and used standard case-control as well as propensity score-matched approaches to assess the
relationship between the heavy metals (Pb, Cd, Se, Hg) and AD risk. Findings from the multivariable
analysis showed the association between BLL and AD risk in tertiles, either in the full population tertile 2
(OR 1.00, 95% CI 0.56-1.79) and tertile 3 (OR 0.87, 95% CI 0.49-1.55) or propensity score-matched
population (tertile 2: OR 1.16, 95% CI 0.55-2.47; and tertile 3: OR 1.12, 95% CI 0.53-2.39), was
imprecise. A cohort study by Horton et al. (2019) used national data from five NHANES cycles (1999-
2008) and followed a large cohort of 8,080 participants from 1999 till December 2014 for AD-related
mortality to examine the longitudinal association between blood Pb and AD mortality. Results from Cox
proportional hazard models adjusted for various confounders and competing risks for AD mortality (death
due to cancer, cardiovascular disease (CVD), cerebrovascular accident [CVA], nephritis, and respiratory
disease) indicated that BLLs of 1.5 and 5 (ig/dL had 1.2 (95% CI = 0.70, 2.1) and 1.4 (95% CI = 0.54,
3.8) times the rate of AD mortality compared with those with a BLL of 0.3 (ig/dL, respectively. The
associations observed for various BLL categories with respect to the reference category of BLL 0.3 (ig/dL
were in a positive direction with increased AD risk for increasing BLL categories; however, the
associations were imprecise. The imprecise effect estimates are likely due to the small number of AD
mortality cases (n = 81), which resulted from AD mortality being determined from the listing of the
immediate cause of death rather than the underlying cause of death. This means the study may be
underpowered, potentially resulting in an unstable effect estimate.
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3.6.4.1.2 Amyotrophic Lateral Sclerosis
Case-control and cohort studies examining the association of BLL and ALS risk or ALS survival
that were included in the 2006 AQCD for Pb or the 2013 Pb ISA produced inconsistent results (Fang et
al.. 2010; Kamel et al.. 2008; Kamel et al.. 2002; Vinceti et al.. 1997). Recent case-control and cohort
studies that assessed biomarkers of Pb before disease development addressed uncertainties related to
temporality and reverse causality identified in previous reviews, thus expanding the support for an
association of BLL with ALS risk and survival.
Strong evidence for the association of Pb and ALS is provided by a recent prospective cohort
study that used data from the National Registry of Veterans in the United States with ALS cases
ascertained between April 2003 and September 2007 (with blood samples collected from January to
September 2007) and followed through the date of death or July 2013 (April 2003-Sep 2007) Fang et al.
(2017). The study was novel in that it assessed ALS mortality and survival and its association with blood
Pb level, and also bone turnover (formation and resorption) biomarkers. The association of ALS survival
time with blood Pb indicated that increased blood Pb was significantly associated with the increased
mortality and thus shorter survival after ALS diagnosis (HR: 1.23 [95% CI: 1.02, 1.49]) in the model
mutually adjusted for bone resorption and formation and other confounding variables. The observation of
the association between BLL and ALS after adjustment for biomarkers of bone turnover reduced
uncertainties related to assessing exposure using Pb concentration in the blood. In another study, Peters et
al. (2020) conducted a nested case-control study within the prospective European Prospective
Investigation into Cancer and Nutrition (EPIC) cohort study. ALS cases were defined to include subjects
with motor neuron disease (ICD10 G12.2) as the underlying cause of death. Pb concentration was
measured in erythrocytes as a marker of ongoing exposure. The associations of Pb in erythrocytes
comparing categories of >56.8 to < 89.0 ng/g and >89.0 ng/g to the reference category (<56.8 ng/g) with
ALS mortality were 1.83 (95% CI: 0.99, 3.35) and 1.89 (95% CI: 0.97, 3.67), respectively.
Additional studies indicating associations of Pb concentration in cerebrospinal fluid (CSF) and Pb
in air also provide some support for an association between Pb exposure and ALS. A case-control study
in Italy used CSF biomarkers for heavy metals including Pb Vinceti et al. (2017). The odds of ALS were
greater in the highest tertile of CSF Pb concentration than in the lowest tertile; however, the effect
estimate was imprecise (OR: 1.39, 95% CI 0.48-4.25). Another case-control study of a large nationally
representative U.S. sample (cases: 26,199 and controls: 78,597) used the U.S. healthcare claims database
from the Symphony Health's Integrated Dataverse (Andrew et al.. 2022). Participants with the first ALS
diagnosis after 6 months of enrollment in the database were included (diagnosis years 2013-2019).
Controls were matched based on age and sex, selected from the Symphony Health network, and also
required to have a minimum of 6 months enrollment in the database. The study did not use the Pb
biomarkers but rather used the airborne contaminants level data for 268 contaminants (including Pb)
obtained from the U.S. EPA National Emissions Inventory (NEI) for 2008 to estimate the past exposure
prior to the ALS onset (2013-2019) at the participants" location of residence. The study used athree-
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phase approach to assess ALS risk nationwide: discovery, validation, and confirmation. First, in the
discovery phase, the study identified major contaminants (out of 268 contaminants) that were
significantly associated with the ALS risk. Second, in the validation phase, the study evaluated various
combinations of contaminants associated with the ALS risk. In the final confirmatory phase, the study
used cohorts only from NH, VT, and OH, incorporating their detailed residential history information to
capture changes in exposure due to residential move prior to ALS diagnosis. The discovery phase
identified 49 airborne contaminants (including Pb) associated with ALS risk. The relationship between
these 49 contaminants and the risk of ALS was further analyzed in the validation cohort, and airborne Pb
and five PCBs were identified associated with an increased risk for ALS (Pb: OR: 1.02, 95% CI: 1.01—
1.03). The confirmatory analysis using NH/VT and OH based cohorts with detailed residential history to
calculate 5-, 10-, and 15-year past exposure prior to diagnosis suggested significant increased risk for
ALS associated with 10-year Pb exposure history when the >75th percentile group was compared with the
<50th percentile group (NH/VT: OR: 2.03, 95% CI: 1.46-2.80, and OH: OR: 1.60, 95% CI: 1.28-1.98) in
a multivariable model. Despite the strong design and larger sample size of the study, the inference
regarding the Pb-ALS risk for this study should be interpreted with caution given the uncertainty
regarding the relationship between the estimated concentration of Pb in the air and Pb concentration in
biomarkers as well as the influence of potential unmeasured confounders.
3.6.4.1.3 Parkinson's Disease
A limited number of case-control studies assessed in the 2013 Pb ISA found positive associations
between bone Pb concentration and PD. A recent study by Paul et al. (2021) examined participants from
two large and independent population-based case-control studies (total n > 2,600)—the System Genomics
of Parkinson's Disease (SGPD), a consortium of three studies from across Australia and New Zealand;
and the Parkinson's Environment and Genes (PEG) study, a population-based study from three
agricultural counties of Central California—to explore the association of cumulative Pb exposure on the
PD risk. The study used novel epigenetic biomarkers of cumulative Pb exposure (i.e., DNA methylation
[DNAm] Pb data in the patella and tibia developed in the NAS cohort). The study analyzed the
relationship between DNAm Pb and PD separately for two cohorts and meta-analyzed the results. The
findings from the multivariable adjusted model suggested that PD risk was strongly associated with the
DNAm biomarker for tibia Pb levels in both cohorts (SGPD cohort: OR: 2.06, 95% CI: 1.66-2.56; PEG
cohort: OR: 1.60, 95% CI: 1.20, 2.15; meta-analyzed results [meta-OR: 1.89, 95% CI: 1.59-2.24]).
3.6.4.1.4 Tremor
A limited number of studies examined the association between BLLs and tremor. The studies
were potentially influenced by reverse causation because inactivity due to disease condition and
subsequent bone resorption can lead to increased BLLs. A recent study used a cohort of men in the NAS
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and assessed the longitudinal relationship of tremor score with bone Pb in the tibia (n = 670, mean: 21.23
(ig/g) and patella (n = 672, mean: 27.98 |ig/g). in addition to BLL (n = 807, mean: 5.01 (ig/dL). Ji et al.
(2015) found that over 8 years of follow-up, neither blood Pb nor bone Pb was associated with the tremor
score. However, among men younger than the median age (68.9 years), the tremor score increased as the
quintile of blood Pb increased (p = 0.03), with men in the highest quintile scoring 0.35 (95% CI: 0.03,
0.67) points higher on the tremor scale than those in the lowest quintile. This pattern was not evident for
bone Pb. The tremor score in the study was based on assessments of drawing capability and not the
clinical diagnosis.
3.6.4.1.5 Motor Function
Several epidemiologic studies examined the association between Pb exposure and decrements in
motor function in adults (Casiens et al.. 2018: Khalil et al.. 2014: Grashow et al.. 2013: Ji et al.. 2013:
Shiue. 2013: Min et al.. 2012). Motor function was assessed using measures of balance, walking speed,
coordination, and strength. Inconsistencies in the results made it difficult to draw conclusions about the
association between Pb exposure and motor function in adults.
Most studies were cross-sectional in design. Results from studies that measured balance were
inconsistent. Among older adults (>50 years) in NHANES 2003-2004, Shiue (2013) observed an inverse
association between urinary Pb and balance disorders defined as self-reported dizziness, difficulty with
balance, or difficulty with falling in the past 12 months. For every log unit increase in urinary Pb (unit not
specified), the likelihood of having a balance disorder decreased (OR = 0.56 (95% CI: 0.38, 0.84]) (Shiue.
2013). Among adults (>40 years) who participated in the NHANES Balance Component, Min et al.
(2012) found that higher levels of blood Pb were generally associated with an increased likelihood of
failing a balance test. Balance dysfunction was evaluated using the Romberg Test of Standing Balance on
Firm and Compliant Support Surfaces, which measured a participant's ability to maintain balance under
various test conditions. Compared with the lowest quintile of blood Pb (<1.2 (.ig/dL). the likelihood of any
balance dysfunction (failing any balance test) increased in the fourth quintile (2.3-3.2 (ig/dL; OR = 5.23
[95% CI: 0.59, 46.43]) and fifth quintile (3.3-48 ^ig/dL; OR = 33.33 [95% CI: 1.94, 573.16]) (Min et al..
2012).
Among older adults (50-85 years) in NHANES, Ji et al. (2013) assessed the relationship between
blood Pb and walking speed. Walking speed was measured by timing a participant's walk for 20 feet at
their usual walking pace. Compared with the lowest quintile of blood Pb (0.2 to <1.2 (.ig/dL). walking
speed decreased with increasing quintiles of blood Pb in women (p-trend = 0.005). In the highest quintile
of blood Pb (3.0 to <53.0 (.ig/dL). the estimated mean walking speed in women was 0.11 feet/second
slower (|3 = -0.11 [95% CI: -0.19, -0.04]). On the contrary, BLL did not appear to be associated with
walking speed in men (Ji et al.. 2013). In MrOS, Khalil et al. (2014) examined the association between
blood Pb and walking speed and strength among older non-Hispanic Caucasian men (>65 years). In this
cross-sectional analysis, blood Pb did not appear to be associated with grip strength, walking speed, or
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narrow-walk pace. Leg extension power (watts) measured with the Nottingham power rig had a negative
association for every log-(ig/dL increase in blood Pb (|3 = -0.03 [95% CI: -1.97, 2.03]). In addition, blood
Pb had a negative association with the participants" ability to stand from a chair without using their arms
(OR per log-(ig/dL increase in blood Pb = 0.97 [95% CI: 0.88, 1.07]) (Khalil et al.. 2014). The reported
associations were very imprecise.
A few cohort studies measured fine motor abilities and hand-eye coordination in adults. In the
HNRS in Germany, Casiens et al. (2018) examined the effect of Pb exposure on fine motor abilities
among male participants who were recruited in 2000-2003 and followed up in 2011-2014. Fine motor
abilities were measured at follow-up (at ages 55-86 years) and included four tasks (tapping, aiming, line
tracing, and steadiness) carried out separately with each hand. Compared with the lowest BLL at baseline
(<5 (ig/dL), the highest BLL (>9 (ig/dL) was positively associated with tapping hits (OR = 1.35 [95% CI:
0.49, 3.70]) and steadiness errors (OR = 1.36 [95% CI: 0.50, 3.66]) but negatively associated with aiming
errors (OR = 0.56 [95% CI: 0.22, 1.42]) and line tracing errors (OR = 0.93 [95% CI: 0.32, 2.74]). The
magnitude of each association increased when using BLLs measured at follow-up, except for tapping hits,
which changed to a negative association (OR = 0.98 [95% CI: 0.82, 1.16]). In general, the ORs were
imprecise (Casiens et al.. 2018). In the Department of Veterans Affairs NAS, Grashow et al. (2013)
examined the association between bone Pb and a coordination task which involved inserting metal pegs
into a grooved pegboard. Bone Pb was measured at the patella and the midtibial shaft and was positively
associated with the grooved pegboard completion time. In other words, the pegboard test took longer to
complete for every 10 (ig/g increase in patella Pb (|3 = 1.97 [95% CI: 0.55, 3.38]) and 10 |ig/g increase in
tibia bone Pb (|3 = 3.11 [95% CI: 1.16. 5.061) (Grashow et al.. 2013).
3.6.4.1.6 Summary
In summary, recent epidemiologic studies found relationships between Pb exposure and some
neurodegenerative disease endpoints among non-occupational cohorts. Similar to the conclusion of the
2013 Pb ISA, the direction and strength of the association is stronger for some endpoints than for others.
In the 2013 Pb ISA, evidence was inconclusive for ALS and AD while a limited number of case-control
studies indicated relationships between higher BLLs in adults and essential tremor, and between higher
bone Pb levels in adults and PD. In the current review, the epidemiologic evidence pertaining to ALS and
PD has strengthened due to the availability of better designed case-control and cohort studies. Recent
studies of Pb exposure and clinically diagnosed AD were also conducted. These studies added to the
previous evidence which generally relied on assessing AD using screening instruments; however, studies
that specifically assessed clinically diagnosed AD or mortality did not provide strong evidence of an
association.
Studies for ALS in this review included one cohort study that examined the relationship between
BLLs and ALS survival among U.S. veterans (Fang et al.. 2017) and a large case-control study of
participants from a healthcare claims dataset examining associations between past airborne Pb exposures
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and ALS risk (Andrew et al.. 2022). The findings from the cohort study by (Fang et al.. 2017) indicate
that increased levels of baseline blood Pb are associated with shorter ALS survival even after mutually
accounting for bone resorption and formation. This finding is in agreement with the results of a study
included in the previous review that suggested positive associations between Pb in blood or bone with
ALS risk (Fang et al.. 2010; Kamel et al.. 2002). An important uncertainty in these studies is the potential
for reverse causality because increased bone turnover in ALS patients could cause higher blood Pb levels,
potentially explaining the observed associations. Fang et al. (2017) addressed this uncertainty by
demonstrating that the association of blood Pb level with ALS survival persisted after adjustment for a
biomarker of bone turnover. Another study investigating ALS risk using a case-control design with a
large sample from a healthcare database and well characterized exposure and outcomes examined whether
previous airborne Pb exposures were related to ALS development. The study found significant positive
associations between Pb exposure and increased ALS risk (Andrew et al.. 2022). specifically for
residential Pb exposure over the past 10 years. The use of estimated airborne exposure without
corresponding measurement of Pb exposure biomarkers is a potential limitation of this study.
In a study of PD, Paul et al. (2021) conducted a case-control analysis using a novel, epigenetic
biomarker to estimate cumulative Pb exposure measured in tibia and patella bone. The study found an
association between DNA methylation (DNAm), as a biomarker of tibia Pb levels, and PD risk. This
empirically-derived DNA methylation signature is potentially a more sensitive predictor of the effect of
Pb exposure on PD than bone Pb concentration (Paul et al.. 2021). With regard to tremor outcomes, no
association with blood and bone Pb biomarkers was observed (Ji et al.. 2015). However, the results
indicated an association of blood Pb and tremor score, particularly in younger men, when analysis was
stratified by age categories. Studies reviewed for Pb exposure and motor function in adults yielded
inconsistent findings.
Studies of the association of Pb exposure with AD in the previous ISA examined cognitive
impairment (e.g., as indicated by MMSE scores, which are used to screen for dementia and AD) rather
than the clinical diagnosis of AD. Recent studies add to the evidence through their examination of clinical
diagnosis of AD and AD mortality in non-occupational cohorts. In a case-control study, Yang et al.
(2018) reported an imprecise positive association between BLL and AD risk. The inability to establish
temporal relationships with this case-control study, given the inclusion of prevalent cases of AD, leads to
uncertainty about potential reverse causality. A recent prospective study that investigated the relationships
between BLLs and AD mortality addressed the temporality of the association (Horton et al.. 2019). The
study calculated HRRs for selected BLLs (i.e., 0.5, 1, 1.5, 2, 3, 5 (ig/dL) compared a reference category of
0.3 (ig/dL. In addition to considering confounders in their model, the authors specified model to consider
the design effect (i.e., accounting for the NHANES survey design utilized by incorporating survey
weights) and competing risks. The study observed a positive association between blood Pb and AD
mortality risk, but the association was imprecise (Horton et al.. 2019). For example, HRR for participants
with BLL of 1.5 1.5 (95% CI = 0.81, 2.9) compared to those with BLL of 0.3 (ig/dl, respectively, after
accounting for design effect.
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3.6.4.2 Toxicological Studies of Neurodegenerative Diseases
Although the evidence was inconclusive overall, a few toxicological studies in the 2013 Pb ISA
suggested that Pb exposure in early life could influence AD-like pathologies (U.S. EPA, 2013). AD
exhibits several neuropathologic hallmarks, such as senile plaques and neurofibrillary tangles (comprised
of A|3 and hyperphosphorylated tau aggregates, respectively), as well as synaptic loss and neuronal death.
Developmental exposure to Pb in rodents, resulting in BLLs >40 (ig/dL, increased A|3 peptides,
hyperphosphorylated tau, and other related endpoints (Li et al„ 2010; Basha et al., 2005). Importantly,
these effects were not found following adult-only exposure. Wu et al. (2008) also demonstrated that 23-
year-old monkeys (Macaco fasciculctris) given Pb from birth to PND 400 had elevated A|3 and amyloid
plaques in their frontal cortex compared with unexposed age-matched controls. BLLs in these animals
ranged from 19 to 26 (ig/dL at PND 400 but had returned to baseline by adulthood. One previous study
performed in transgenic superoxide dismutase 1 (SOD1) mice (a model of ALS) found that adolescent
exposure to Pb reduced astrocyte reactivity and extended the survival time but had no significant effects
on the onset of disease in this model (Barbeito et al., 2010). Tavakoli-Nezhad et al. (2001), reviewed in
the 2006 AQCD, demonstrated Pb-induced decreases in dopaminergic cell activity in the substantia nigra,
which is associated with PD. No recent PECOS-relevant studies have investigated the effects of Pb
exposure on ALS-relevant endpoint or endpoints related to essential tremor.
The potential relationship between Pb exposure and AD has been further explored in recent
literature (Table 3- 17T). Of the two major A|3 isoforms (A|340 and A|342), the less predominant isoform,
A|342, is typically considered more prone to aggregation (Xiao et al„ 2015). A low A|342/A|340 ratio in
plasma and CSF has also been shown to indicate an increased risk for AD (Graff-Radford et al., 2007).
(Zhou et al., 2018) found increased expression of A|342 in the cerebral cortex and hippocampus of
Sprague Dawley rats following Pb exposure during adolescence. Gestational and lactational Pb exposure
resulting in lower BLLs, between 4-10 (ig/dL, also significantly increased A|340 in the cerebral cortex of
Kunming mice (Li et al., 2016c). Utilizing a transgenic mouse model (Tg-SwDI), Gu et al. (2012)
reported that adolescent Pb exposure significantly increased both A|340 and A|342 in the cerebral cortex,
hippocampus, and CSF; however, the ratio of A|342/A|340 was not significantly different. Pb-treated
animals in this study also had increased amyloid plaque formation, which was co-localized with brain Pb
deposits. Increases in the expression of amyloid precursor protein (APP) and beta-secretase 1 (BACE1;
the enzyme that cleaves APP into A|3) have been demonstrated in some recent studies (Wu et al., 2020b;
Zhou et al„ 2018; Sun et al., 2014), but not all (Gu et al„ 2012).
Changes in the expression of both total tau (t-tau) and p-tau are considered biomarkers of AD in
humans. One recent study demonstrated that developmental Pb exposure (GD 0-PND 21; resulting in
BLLs of 7 (ig/dL) increased t-tau and p-tau in the cerebral cortex and cerebellum but not the hippocampus
of juvenile Wistar rats (Gassk et al., 2016b). These changes coincided with some evidence of
enhanced activity of two tau kinases (glycogen synthase kinase-3|3 and cyclin-dependent kinase 5
[CDK5]) in relevant brain regions. Another study found that postnatal Pb exposure in Wistar rats caused
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significant changes in t-tau, p-tau, and related phosphatases in the hippocampus, but these effects were
transient and inconsistent at the time points tested (PND 21 and PND 30) (Rahman et al.. 2012b). Wu et
al. (2020b) also found no significant increases in hippocampal p-tau in Pb-exposed C57B1/6 mice at 4
months old; however, increases in hippocampal p-tau became apparent at 13 months old and persisted
until 16 months old. In the prefrontal cortex, p-tau was significantly elevated above age-matched control
values at 4 month and 13 months, but not 16 months. Importantly, BLLs in these rodents at 4 months
(when exposure was terminated) were nearly 60 (ig/dL and did not decrease to PECOS-relevant values
until 16 months (28 (.ig/dL). One additional study Zhang et al. (2012) reported that 8 weeks of Pb
exposure increased p-tau in the hippocampus at BLLs as low as 10 (ig/dL. Notably, this study also
reported increased alpha-synuclein in the hippocampus. Alpha-synuclein is a major constituent of Lewy
bodies, which are a neuropathologic hallmark of PD (Baba et al.. 1998).
In the 1980s, Rice (1990) established a cohort of monkeys {Macaca fascicidaris) exposed to Pb in
the first 400 days of life (resulting in BLLs between 19-26 (ig/dL) and terminated at 23 years old. At the
time of termination, BLLs had returned to control levels. Using tissues from these animals, Wu et al.
(2008) found increases in the protein expression of A|3 and APP, as well as increases in the gene
expression of APP and specificity protein 1 (Spl, a transcriptional regulator of APP and tau), reported in
the previous ISA. Recently, Bihaqi and Zawia (2013) extended these findings by analyzing the cerebral
cortex tissue for changes in tau-related endpoints. Compared with age-matched controls, Pb-exposed
animals had significantly increased expression of t-tau, p-tau, and CDK5 (a tau kinase). These findings
were further supported by neuropathological changes (i.e., increases in p-tau immunoreactivity and
deposits). This study also found that mRNA levels of tau, CDK5, Spl, and Sp3 were significantly
increased.
Recent studies have also measured endpoints outside of those related to A|3 and tau.
Dysregulation of lipid pathways has been implicated in AD and neurodegenerative disorders (Di Paolo
and Kim. 2011). Zhou et al. (2018) found that Pb exposure decreased total and free cholesterol levels via
dysregulation of cholesterol metabolism in the cerebral cortex and hippocampus. Feng et al. (2019) found
that lifetime exposure to Pb significantly decreased neuronal density in the cerebral cortex at PNW 70.
This change was accompanied by a decrease in overall brain volume. In addition, several studies that
reported on AD-related neuromolecular changes also reported significant impairment of learning and
memory assessed via the Morris water maze (Wu et al.. 2020b; Li et al.. 2016c; Gu et al.. 2012; Rahman
et al.. 2012b). However, these results cannot be definitively attributed to AD-related neurobehavioral
changes due to the well-known effects of Pb on cognitive function, which are unrelated to AD.
Deficiencies in balance, walking speed, coordination, and strength can also arise from insults to
the motor system in adulthood. Recent toxicological studies exposed mature rodents to Pb and
investigated the effects on motor function. Typical rotarod tests compare the latency to fall for subjects
placed on a rotating rod. Falling off more quickly indicates decreased coordination or balance. Locomotor
activity tests (e.g., measurements of distance traveled, counts of square crossings) can detect gross motor
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problems as well; however other influences on behavior may factor into differences in the amount of
movement. Fine motor forelimb grip strength can be determined in rodents by pulling subjects holding
onto a measurement-taking tension bar. Mansouri et al. (2012) subjected Wistar rats to Pb acetate for 30
days and observed hyperactivity in open-field tests in males but not females on the final day (PND 100).
They also observed no effect on rotarod performance in both males and females on PND 100. However,
in a subsequent study, Mansouri et al. (2013) found that long-term exposure to Pb acetate in drinking
water (155-159 days starting at PND 55-60) resulted in substandard rotarod performance (7.5 months
old) for male Wistar rats, whereas the exposure had no effect on performance of female rats. Similarly, in
a long-term exposure study by Singh et al. (2019). male Wistar rats exposed daily to Pb acetate by oral
gavage from 3 months to 6 months of age performed worse in rotarod and grip strength tests compared
with their saline-treated counterparts. Singh et al. (2019) also found decreased activity in 6-month-old Pb-
treated rats. Al-Qahtani et al. (2022) observed a decrease in locomotor activity in 15-week-old male mice
after a 6-week Pb treatment period.
Summary
In summary, recent studies have significantly expanded the toxicological literature base
established in the last Pb ISA for AD. Significant increases in A|340 and A|342 following Pb exposure
were consistently detected in the cerebral cortex, hippocampus, and CSF in multiple studies at BLLs of 4-
30 (ig/dL. Pb-induced amyloid plaque formation was also reported in a transgenic mouse model of AD
(Tg-SwDI) (Gu et al.. 2012). Aged cynomolgus monkeys (23 years old), exposed to Pb during infancy,
had both amyloid plaques and tau deposits in their cerebral cortex; however, these findings are limited
somewhat by the small sample size (Bihaqi and Zawia. 2013; Wu et al.. 2008). In rodents, mean BLLs
<10 (ig/dL were shown to increase the expression and phosphorylation of tau in multiple brain regions,
but this effect was not entirely consistent between studies. Overall, recent studies have primarily focused
on exposure paradigms beginning during development, but one study demonstrated effects from an
exposure beginning in early adulthood (8 weeks) (Gu et al.. 2012). One study reported that the PD-related
protein, alpha-synuclein, was increased in the hippocampus in Pb-treated male rats (Zhang et al.. 2012).
Recent studies have not expanded on previous findings on ALS-related endpoints or contributed evidence
related to essential tremor.
3.6.4.3 Relevant Issues for Interpreting the Evidence Base
3.6.4.3.1 Concentration-Response Function
The shape of the C-R function was not examined in the studies of the association of Pb
biomarkers with neurodegenerative diseases in adults in the past review. The majority of studies in the
current review also did not explore the shape of the C-R function in the Pb-neurodegenerative disease
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associations. Horton et al. (2019) explored the relationship between BLL and AD mortality using Cox
regression models that incorporated design effect or competing risks. The study found an increase in the
hazard rate ratio (HRR) by 14%-30% with each unit increase in BLL (see Figure 3-15 below). Given the
small number of AD mortality events in the study population, the effect estimates were imprecise and had
larger CIs. The authors also performed categorical analysis to explore blood Pb-AD mortality association.
Results from Cox proportional hazard models adjusted for various confounders and competing risks for
AD mortality (death due to cancer, CVD, CVA, nephritis, and respiratory disease) indicated that
participants in the 1.5 and 5 (ig/dL BLL categories had 1.2 (95% CI = 0.70, 2.1) and 1.4 (95% CI = 0.54,
3.8) times the rate of AD mortality compared with those with a blood Pb reference of 0.3 (ig/dL,
respectively. The associations observed for various BLL categories with respect to the reference category
of BLL 0.3 (ig/dL were imprecise.
BLL = blood lead level; HRR = hazard rate ratio; LCI = lower confidence interval.
Source: Reproduced with permission from Horton et al. (20191.
Figure 3-15 Hazard rate ratios for Alzheimer's disease mortality by blood Pb
level including the lower 95% confidence interval.
1.5 2 2.5 3 3.5 4 4.5 5
BLL (ng/dL)
Design effect
— — — LCI-Design effect
Null line
Competing risk
— — — LCI-Competing risk
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3.6.4.3.2 Potentially At-Risk Populations
Genetics
Several studies in the 2013 Pb review evaluating the association between Pb and MMSE (a
marker for AD) and effect modification by genetic variants provided support for effect modification of
the association by the ALAD genotype (details on Potentially At-Risk Populations in the Cognitive
Function Section: 3.6.1.3.2). A study by (Fang et al.. 2010) examining the association of BLL and ALS
risk and effect modification by the ALAD genotype suggested a significant Pb-ALS association among
ALAD1-1 carriers but a weaker and imprecise association among ALAD2 carriers. Tests to identify an
interaction between Pb and the ALAD genotype in the Pb-ALS association suggested no significant
difference in association between ALAD1-1 versus ALAD2 carriers, however (p = 0.32).
In the current Pb review, (Ji et al.. 2015) considered both bone and blood Pb biomarkers among
the NAS cohort and performed stratified analysis by ALAD gene (ALAD-2 carriers or non-carriers).
They found no effect modification by the ALAD genotype for the association between Pb biomarkers and
elevated tremor.
Age and Sex
A few studies in the current review explored the effect modification of the Pb-neurodegenerative
disease associations by age or sex. (Ji et al.. 2015) examined the associations of bone and blood Pb
biomarkers with tremor among the NAS cohort. The results suggested that among younger cohorts (i.e.,
below the median age of 68.9 years), the tremor score increased significantly with increasing quintile of
blood Pb (p = 0.03), and those in the highest quintile scored 0.35 (95% CI: 0.03, 0.67) points higher than
those in the lowest quintile. This pattern was not apparent when bone Pb biomarkers were used. Similarly,
(Paul et al.. 2021) performed stratified analysis of the DNAm estimated tibia and patella Pb
concentrations and PD risk by sex and found a significant association when tibia Pb concentration was
used. The magnitude of risk was higher for men in the SGPD cohort (OR and 95% CI: men: 2.48 [1.86,
3.34]; women: 1.67 [1.21, 2.33]), and the risk was higher for women in the PEG cohort (OR and 95% CI:
men: 1.49 [1.02, 2.20]; women: 1.81 [1.17, 2.85]).
3.6.4.4 Summary and Causality Determination: Neurodegenerative Diseases
The 2013 Pb ISA (U.S. EPA, 2013) concluded that the available evidence was "inadequate to
determine that a causal relationship exists between Pb exposure and neurodegenerative diseases in
adults." This conclusion was based on a limited number of studies that examined the association of blood
Pb or bone Pb levels with essential tremor, PD, ALS, and AD. These studies were not sufficient to reach a
conclusion regarding the presence or absence of an effect due largely to the potential for reverse causation
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(i.e., reduced physical activity among cases resulting in greater bone turnover and higher BLLs), and
limited consideration for potential confounding factors. Limited studies in monkeys and rodents found
that developmental Pb exposure induced pathologies that underlie AD, and rodent studies suggested
neurophysiologic characteristics and changes related to ALS and PD. Recent epidemiologic studies
expanded the evidence base indicating associations between Pb exposure and some neurodegenerative
diseases among non-occupational cohorts. The strongest evidence in the current review includes well-
designed case-control and cohort studies of ALS and PD, whereas findings from recent epidemiologic
studies of clinically diagnosed AD or AD mortality did not provide strong evidence to address
uncertainties in the body of evidence. Findings from recent toxicological studies, however, add to the
evidence suggesting that developmental exposure to Pb increases the expression of proteins related to
AD, including A|3, tau and p-tau at lower BLLs than the values investigated in the previous ISA (<10
(ig/dL). Alterations in neuropathologic hallmarks of AD in older monkeys were also demonstrated
following developmental Pb exposure.
Studies for ALS in this review included a well-designed cohort study that examined the
relationship of the blood Pb biomarker and ALS survival after ALS diagnosis among U.S. veterans (Fang
et al.. 2017) and a large case-control study that examined associations between past airborne Pb exposures
and ALS risk in participants from a healthcare claims dataset (Andrew et al.. 2022). The findings from the
cohort study by (Fang et al.. 2017) suggested that increased levels of past blood Pb prior to mortality
follow-up were associated with shorter ALS survival even after mutually accounting for bone resorption
and bone formation, thus reducing the uncertainty due to reverse causality. (Andrew et al.. 2022) found
significant positive associations between airborne Pb exposure and increased ALS risk, specifically for
residential Pb exposure over the past 10 years. For PD outcomes, only one case-control study investigated
the relationship between the risk of PD and epigenetic biomarkers by quantifying DNAm tibia and patella
Pb concentrations as cumulative Pb exposure. The authors found an association between DNAm tibia Pb
levels and PD risk (Paul et al.. 2021). This empirically-derived DNA methylation signature is potentially
a more sensitive predictor of the effect of Pb exposure on PD than bone Pb concentration.
Toxicological studies investigating potential associations between Pb exposure and ALS or PD
remain limited. In one study reviewed in the previous ISA, Pb exposure was found to induce
neurophysiologic changes in a rodent model of ALS. Neurophysiologic characteristics of PD, such as
decreased activity of dopaminergic neurons in the substantia nigra and increased expression of
hippocampal alpha-synuclein, have also been demonstrated following Pb exposure.
The bulk of the epidemiologic evidence in the 2013 Pb ISA drawn upon to evaluate the
association of Pb exposure AD focused on cognitive impairment identified using dementia screening
instruments such as the MMSE. Recent studies that examined the association of blood Pb biomarkers
with clinical endpoints of AD risk (Yang et al.. 2018) or AD mortality (Horton et al.. 2019) in non-
occupational cohorts add to the evidence. The association between blood Pb exposure and AD risk
observed in Yang et al. (2018) was imprecise. Uncertainties related to potential reverse causality and
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timing of exposure were not addressed in this study. (Horton et al.. 2019) addressed concerns raised for
the case-control study design used in Yang et al. (2018) but also observed imprecise relationships
between blood Pb and AD mortality risk, (Horton et al.. 2019). For tremor outcomes, no significant
association was observed between blood Pb and the tremor score when all study participants were
analyzed; however, an increase in the tremor score was observed for increasing BLL among younger
participants (Ji et al.. 2015). Studies reviewed for Pb exposure and motor function in adults also provided
inconsistent findings. Overall, the relationships of Pb biomarkers with AD risk, tremor, or motor function
are inconclusive. In contrast to the inconclusive epidemiologic study findings on AD, a large number of
recent toxicological studies add to the evidence which suggests that developmental exposure to Pb
increases the expression of proteins related to AD, including A|3, tau, and p-tau, at lower BLLs than the
values investigated in the previous ISA (<10 (.ig/dL). In older monkeys (23 years), the neuropathologic
hallmarks of AD (i.e., amyloid plaques and tau deposits) were also demonstrated following
developmental Pb exposure.
In summary, the evidence from epidemiologic and experimental animal studies is suggestive
of, but not sufficient to infer, a causal relationship between Pb exposure and neurodegenerative
diseases. This determination reflects a strengthening of the evidence since the 2013 Pb ISA, which found
that the evidence was "inadequate." Recent epidemiologic studies of varying quality strengthen the
evidence for the association of Pb exposure with ALS and PD, and reduce uncertainty related to the
potential for reverse causality by better establishing the temporal association between Pb exposure and
these clinical endpoints. Although recent epidemiologic studies of clinically diagnosed AD and AD
mortality add to the evidence, findings from these studies do not substantially strengthen the evidence
overall. In contrast to the epidemiologic evidence, multiple recent toxicological studies add to the
evidence indicating that developmental exposure to Pb increases the expression of proteins related to AD,
including A|3, tau, and p-tau at lower BLLs than the values investigated in the previous ISA (<10 (.ig/dL).
Alterations in neuropathologic hallmarks of AD in older monkeys were also demonstrated following
developmental Pb exposure. However, toxicological studies investigating potential associations between
Pb exposure and ALS or PD remain limited.
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Table 3-12 Summary of evidence that is suggestive of, but not sufficient to infer, a causal relationship
between Pb exposure and neurodegenerative diseases in adults
Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker
Levels Associated
with Effects0
ALS and Parkinson's Disease
At least one high-quality
prospective cohort or case-control
study finds associations with ALS
and PD.
A prospective analysis of U.S. veterans found that higher baseline BLL was Fang et al.
associated with increased mortality / shorter survival after ALS diagnosis. The (2017)
association persisted after controlling for confounders including a biomarker of bone
turnover (formation and reabsorption), thus addressing the issue of reverse causality.
Support from recent case-control studies using novel exposure metrics that found Paul et al.
associations between higher estimated air Pb exposure and ALS risk and between a (2021)
DNAm biomarker of tibia Pb and PD.
Limited number of studies address Well-designed case-control studies and prospective studies of ALS and PD assessed
uncertainty due to temporality and exposure prior to disease development and accounted for increased bone turnover
reverse causation. resulting from the disease state.
Alzheimer's Disease
Coherence for AD provided by
consistent evidence in animals
with relevant exposures.
Amyloid plaques and/or tau deposits in transgenic rodents and aged monkeys
following Pb exposure.
Bihaai and Peak BLLs: 19-30
Zawia (2013) pg/dL
Increased expression of A(3, tau, and other AD-related proteins across multiple brain Wu et al.
regions in rodents. (2008)
Gu et al.
(2012)
Section 3.6.4.2
Peak BLLs: 4-58
pg/dL
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Rationale for Causality
Determination3
Key Evidence13
Pb Biomarker
References'3 Levels Associated
with Effects0
Evidence describes biologically
plausible pathways.
Evidence suggests that exposure to Pb results in neuronal cell death associated with Section 3.3
oxidative stress, neuroinflammation and altered energy metabolism, all of which may
underlie general neurodegenerative processes.
AD = Alzheimer's disease; ALS = amyotrophic lateral sclerosis; BLL = blood lead level; DNAm = DNA methylation; Pb = lead; PD =Parkinson's disease.
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs CU.S. EPA. 20151.
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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3.7
Evidence Inventories - Data Tables to Summarize Study Details
Table 3-1E Epidemiologic studies of Pb exposure and overt nervous system toxicity
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Yuan et al. (2006)
Cincinnati, OH
CLS
n: 24
Blood
Age at measurement:
BLL from 3 to 78 mo
averaged
Mean: 14.18
Range: 4.77-31.06
[jg/dL
MRI (subject asked to generate birth weight and
verbs to activate language with marijuana use;
bilateral finger tapping) consideration of IQ,
sex, SES, gestational
age
Increasing BLL
associated with
decreased brain
activation in the left
frontal gyrus and left
middle temporal gyrus,
regions (semantic
language function)
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tReuben et al. (2020)
Dunedin
New Zealand
1972-2019
Cohort
Dunedin Study
n: 512
Whole blood Pb (pg/dL)
was measured via
GFAAS
Age at measurement:
11 yr
Mean (SD): 10.99 (4.63)
|jg/dL
Max: 31 |jg/dL
Cortical thickness, cortical
surface area, hippocampal
volume, WMH volumFe,
BrainAGE index
High resolution images
showing cortical thickness,
cortical surface area, bilateral
hippocampal volume, WMH,
and FA were produced using
T1-weighted, fluid-attenuated
inversion recovery and
diffusion-weighted sequences
with a Siemens Skyra 3T
scanner with 64-channel head
and neck coil. BrainAGE index
was calculated as a composite
measure of all measured
indices. Outcomes were
assessed at 45 yr of age.
Sex, maternal IQ,
childhood :
Betas
BrainAGE Index: 0.03
(0.00, 0.06)
Hippocampal Volume:
0.00 (-0.01, 0.00))
Cortical Surface Area:
-0.05 (-0.09, 0.00)
Cortical Thickness
(mm):
0.00 (0.00, 0.00)
WMH:
0.00 (0.00, 0.01)
Age at outcome:
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Cecil (2011)
Cincinnati, Ohio, Cincinnati
Children's Hospital Medical
Center
United States
Enrollment mothers): 1979-
1984. Follow-up: Birth to 24 y
Cohort
CLS
n: 159
Whole blood Pb
measured using anodic
stripping voltammetry
Age at measurement:
3-78 mo old
Pb prenatally and at
intervals to age 17 yr;
imaging 19-24 yr
Mean (SD) mcg/dL:
prenatal 8.3 (3.7), 3-12
mo 10.6 (5.1).
(Reported in Dietrich et
al. 1993)
MRI brain assessments of 4
types: volumetric (morphology),
spectroscopy (chemical
concentrations), diffusivity
(organization), and functionality
(activation related to tasks).
Brain MRI/fMRI measures of
four types. (1) Volume of gray
matter. (2) Spectroscopy -
metabolites linked to neuronal
function and myelin
architecture: N-acetyl
aspartate, creatine and
phosphocreatine,
phosphocholines and
glycerolphosphocholine, myo-
inositol, glutamate and
glutamine. (3) Diffusivity in
white matter regions reflecting
axonal and myelin effects: FA;
mean, axial and radial
diffusivity. (4) Functionality:
activation related to task
performance.
Age at outcome:
19-24 yr
Varied by outcome;
included age at imaging
and birth weight.
Reported a negative
association between
childhood BLL and gray
matter volume in
several regions: medial
and superior frontal
gyri, inferior parietal
lobule, cerebellar
hemispheres
Reported an
association between
higher childhood BLL
and lower metabolite
concentrations in
several regions: white
matter, left basal
ganglia, left cerebellar
hemisphere, vermis
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tBeckwith et al. (2021)
Cincinnati, OH
United States
in utero to up to 33 yr of age
Cohort
CLS
n: 123
Pb was measured in
whole blood at 10 d, on
a quarterly basis up to
60 mo, and monthly at
66, 72, and 78 mo.
Samples were collected
mainly by venipuncture,
but occasionally by heel
or finger stick. Pb
concentrations were
quantified using anodic
stripping voltammetry
Age at measurement:
0-78 mo
10 days, on a quarterly
basis up to 60 mo, and
monthly at 66, 72, and
78 mo
MRI brain volumetrics of white
and gray matter, focusing on
regions involved in cognitive
and emotional function.
MRI scans (Voxel based
morphometry) were used to
examine spatial differences in
regional gray and white matter
volumes in adulthood (mean
age 26.8 yr) associated with
childhood blood Pb
concentrations at 78 mo.
Age at outcome:
18-33 yr
Age at time of imaging,
birth weight, total
intracranial volume.
BLLs were associated
with MRI-derived
decreases in white and
gray matter volumes in
the frontal, parietal, and
temporal lobes.
Decreased gray matter
volume in brain regions
responsible for
cognition and emotional
regulation associated
with criminal arrests
Mean (SD) blood Pb at
78 mo: 7.82 pg/dL (4.2)
Max: 24.75 pg/dL
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tLamoureux-Tremblav et al.
(2021)
Nunavik, Northern Quebec
Canada
Cohort
NCDS
n: 71
Blood
Cord and concurrent
blood; GFAAS with
Zeeman background
correction
Age at measurement:
16-22
Cord blood: median
3.73 |jg/dL, mean 4.56
pg/dL. Adolescent
blood: median 1.52
|jg/dL, mean 1.78 |jg/dL
Max: 17.81 pg/dL
Activation of the human neural
fear circuitry
fMRI analysis of brain
activation in response to a
validated fear conditioning and
extinction stimulus test
Age at outcome:
16-22 yr
Sex, age, SES, and
alcohol/drug
consumption.
Higher differential
activation in the right
dorsolateral prefrontal
cortex in association
with higher postnatal
BLL.
tEthieret al. (2012)
Nunavik, Quebec
Canada
11 yr
Cohort
Prospective 11 yr Blood, Maternal Blood Neurological
study of Inuit
children from
Nunavik
n: 149
Concurrent venous
blood; GFAAS with
Zeeman background
correction (Perkin Elmer
model ZL 4100).
Age at measurement:
Pre-natal and 11 yr
At birth mean: 4.6
pg/dL, SD: 3.1; At 11 yr
mean: 2.6 pg/dL; SD:
2.3
Achromatic pattern-reversal
VEPs with different visual
contrast levels were
administered, using generated
vertical sinusoidal gratings with
a spatial frequency of 2.5
cycles per degree. Children
viewed stimuli binocularly and
were instructed to fixate on a
small red dot. Pattern-reversal
VEPs were recorded from the
scalp over the visual cortex at
Oz derivation according to the
International 10-20 system
Age at outcome:
10-13 yr
Analysis of variance
models controlled for
current Se, cord Se,
and gender.
Betas
N150 Latency
95% Contrast Level:
0.056 (0.099, 0.014)
*Note- 95% CIs were
converted from author
reported p-values
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tKimetal. (2018a)
n: 150
Seoul
Korea
Enrollment 2010-
Case-Control
¦2015
Blood
Cortical Thickness
Blood Pb was measured Cortical thickness of brain
using atomic absorption regions was ascertained via
Age, intracranial
volume, gender.
spectrometer graphite
furnace
Age at measurement:
6-17 years
Mean (SD) - Cases: 1.3
(0.6) |jg/dL; Controls:
1.5 (0.7) |jg/dL
whole-brain structural MRI.
Age at outcome:
6-17 yr
An interaction between
DRD2 and BLL on the
cortical thickness of the
frontal lobe in the
ADHD group, and a
brain-behavior
correlation between
cortical thickness and
the ADHD-RS
inattention score was
observed.
BLL = blood lead level; BrainAGE = Brain Age Gap Estimation; CI = confidence interval; CLS = Cincinnati lead study; d = day(s); FA = fractional anisotropy; fMRI = functional
magnetic resonance imaging; GFAAS = graphite furnace atomic absorption spectrometry; IQ = intelligence quotient; mo = month(s); MRI = magnetic resonance imaging; Pb = lead;
SD = standard deviation; Se = selenium; SES = socioeconomic status; VEP = visual evoked potential; WMH = white matter hyperintensities; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResults are not standardized (e.g., BLL distribution data needed to calculate the standardized estimate was not reported or categorical data was analyzed).
tStudies published since the 2013 Integrated Science Assessment for Lead.
Table 3-1T Animal toxicological studies of Pb exposure and brain function
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Graham et al. (2011)
Rat (Sprague Dawley) PND 4 to PND 28, Oral, gavage PND 29:
Control (vehicle), M/F, n every other day
= 4-8 0.289 pg/dL for Control
1 mg/kg, M/F, n = 4-8
10 mg/kg, M/F, n = 4-E
3.27 pg/dL for 1 mg/kg
12.6 pg/dL for 10 mg/kg
PND 11, 19, 29: Neurotransmitter
3-233
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Study Species (Stock/Strain), TimingoJ Exposurc BLL as Reported (Mg/dL) Endpoints Examined
Liu et al. (2012) Rat (Sprague Dawley) PND24toPND80 Oral, drinking PND21: PND 56: Electrophysiology
Control (tap water), M, n water
= 20 15 |jg/L (1.5 |jg/dL) for Control
100 ppm, M, n = 20 45 |jg/L (4.5 |jg/dL) for 100
ppm
PND 28:
14 |jg/L (1.4 |jg/dL) for Control
94 |jg/L (9.4 pg/dL) for 100
ppm
PND 35:
16 |jg/L (1.6 pg/dL) for Control
103 |jg/L (10.3 pg/dL) for 100
ppm
PND 42:
13 |jg/L (1.3 pg/dL) for Control
94 |jg/L (9.4 pg/dL) for 100
ppm
PND 49:
14 |jg/L (1.4 pg/dL) for Control
98 |jg/L (9.8 pg/dL) for 100
ppm
3-234
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Corv-Slechta et al. (2012) Rat (Long-Evans) GD-60 to 10 mo Oral, drinking PND 5-6
Control (tap water), M, n water
= 12 Oral,
lactation
50 ppm Pb, M, n = 12 In utero
10-11 mo: Neurotransmitter
<5 pg/dL for Control
12.5 pg/dL for 50 ppm
2.5 mo:
<5 pg/dL for Control
6.43 pg/dL for 50 ppm
10 mo:
<5 pg/dL for Control
8.98 pg/dL for 50 ppm
Weston et al. (2014)
Rat (Long-Evans) GD -60 to PND 21
Oral,
PND 5-6 - Males:
PND 60: Neurotransmitter
Control (tap water), M/F,
lactation
n = 18-22 (9-11/9-11)
In utero
0.76 pg/dL for Control
50 ppm, M/F, n = 18-22
15.7 pg/dL for 50 ppm
(9-11/9-11)
PND 5-6 Females:
0.82 pg/dL for Control
14.7 pg/dL for 50 ppm
3-235
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Han et al. (2014)
Rat (Wistar) PW group: PND 21 Oral, drinking PND21:
Control (tap water), M, n to PND 42 water
= 8 Oral,
ME group: GD -21 lactation
2 mM - postweaning to PND 20 In utero
(PW), M, n = 8
PND 68: Histopathology
7.36 |jg/L (0.74 pg/dL) for
Control
2 mM - ME, M, n = 8
NR for 2 mM - PW
146.6 pg/L (14.7 pg/dL) for 2
mM-ME
PND 63:
9.22 pg/L (0.92 pg/dL) for
Control
147.9 pg/L (14.8 pg/dL) for 2
mM-PW
46.13 pg/L (4.6 pg/dL) for 2
mM-ME
Barkur and Bairv (2015a) Rat (Wistar)
Pregestation
Control (untreated), M/F, exposure - GD -30
n = 9
to GD 0
Lactation only
0.2% solution -
Pregestation, M/F, n = 9 exposure - PND 0
to PND 22
0.2% solution -
Lactation, M/F, n = 9
0.2% solution -
Gestation, M/F, n = 9
0.2% solution -
Gestation and Lactation,
M/F, n = 9
Gestational
exposure - GD 0 to
GD 20
Gestation and
Lactation exposure
-GD Oto PND 22
In utero PND 22:
0.19 pg/dL for Control
3.04 pg/dL for 0.2%
Pregestation
5.26 pg/dL for 0.2% Gestation
26.8 pg/dL for 0.2% Lactation
31.9 pg/dL for 0.2% Gestation
and Lactation
PND 30: Brain Weight
3-236
-------
Study Species (Stock/Strain), TimingoJ Exposurc BLL as Reported (Mg/dL) Endpoints Examined
Bashaetal. (2014)
Rat (Not Specified)
PND 1 to PND21
Oral,
PND 45:
PND 45, 4 mo, 12 mo, 18 mo:
Control (deionized
lactation
Neurotransmitter Analysis
water), M, n = 6
0.42 |jg/dL for Control
0.2% solution, M, n = 6
49.5 |jg/dL for 0.2% solution
4 mo:
0.56 |jg/dL for Control
14.4 |jg/dL for 0.2% solution
12 mo:
0.46 |jg/dL for Control
6.96 |jg/dL for 0.2% solution
18 mo:
0.12 |jg/dL for Control
11.2 |jg/dL for 0.2% solution
Rahman et al. (2012b)
Rat (Wistar)
PND 1 to PND 30
Oral, drinking
PND 21:
PND 21, 30: Brain Weight,
Control (tap water), M/F,
water
Histopathology
n = 4-10
Oral,
1.4 |jg/dL for Control
lactation
0.2% solution, M/F, n =
12.1 |jg/dL for 0.2% solution
4-10
PND 30:
1.2 |jg/dL for Control
12.8 |jg/dL for 0.2% solution
3-237
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Nam et al. (2019b)
Rat (Sprague Dawley) GD 0 to 22
Control (not specified),
M/F, n = 12
0.2 % Solution, M/F, n =
12
In utero PND21:
0.64 pg/dL for Control
17.30 pg/dL for 0.2% solution
PND21: Histopathology
Saleh et al. (2018)
Rat (Sprague Dawley) GD 1 to GD 20
Control (deionized
water), M/F, n = 8 litters
160 ppm, M/F, n = 8
litters
In utero Maternal Blood Pb GD 20:
5.1 pg/dL for Control
27.7 pg/dL for 160 ppm
GD 20: Brain Weight, Histopathology
Menq et al. (2016)
Rat (Sprague Dawley) PND 0 to PND 21
Control (deionized
water), M/F, n = 7
300 ppm, M/F, n = 7
Oral,
lactation
PND 35:
7.61 |jg/L (0.76 pg/dL) for
Control
84.3 pg/L (8.43 pg/dL) for 300
ppm
NR: Histopathology
Amos-Kroohs et al.
(2016)
Rat (Sprague Dawley) PND 4 to PND 28
Control (sodium
acetate), M/F, n = 16
(8/8) per time point
1 mg/kg Pb, M/F, n = 16
(8/8) per time point
10 mg/kg Pb, M/F, n =
16 (8/8) per time point
Oral, gavage PND 29:
1.27 pg/dL for Control
2.76 pg/dL for 1 mg/kg
9.07 pg/dL for 10 mg/kg
PND 29: Neurotransmitter
3-238
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Rahman et al. (2018)
Rat (Wistar)
Control (tap water), M/F,
n = 20
0.2% solution, M/F, n =
20
PND 1 to PND 21 Oral, drinking PND 21:
PND 21, 30: Brain Weight
water
Oral,
lactation
2.2 pg/dL for Control
12.4 pg/dL for 0.2% solution
PND 30:
3.3 pg/dL for Control
22.7 pg/dL for 0.2% solution
Baranowska-Bosiacka et Rat (Wistar) GD 0 to PND 21 Oral, PND 28:
al. (2017) Control (distilled water), lactation
M/F, n = 8 In utero 0.05 pg/dL for Control
0.1% solution, M/F, n =
6.90 pg/dL for 0.1 % solution
PND 28: Histopathology
Baranowska-Bosiacka et Rat (Wistar)
al. (2013) Control (distilled water),
M/F, n = 36 (17/19)
0.1% solution, M/F, n =
36 (18/18)
GD Oto PND 21
Oral,
lactation
In utero
PND 28:
0.93 pg/dL for Control
6.86 pg/dL for 0.1 % solution
PND 28: Histopathology
Wang et al. (2013)
Rat (Sprague Dawley) GD 0 to PND 1,
Control (untreated), M/F, PND 1 to PND 21,
n = 6 PND 21 to 42
0.2% Pb (w/v), M/F, n =
6 - Gestational
Exposure
0.2% Pb (w/v), M/F, n =
6 - Lactational Exposure
0.2% Pb (w/v), M/F, n =
6 - Ablactational
Exposure
Oral, drinking PND 72:
water
Oral,
lactation
In utero
PND 72: Brain Weight
34.99 |jg/L (3.5 pg/dL) for
Control
35.78 pg/L (3.58 pg/dL) for 0.2
% solution Gestational
65.97 pg/L (6.60 pg/dL) for
0.2% solution Lactational
110.67 pg/L (11.07 pg/dL) for
0.2% solution Ablactational
3-239
-------
Study Species (Stock/Strain), TimingoJ Exposurc BLL as Reported (Mg/dL) Endpoints Examined
Wang et al. (2016) Rat (Sprague Dawley) PND24toPND56 Oral, drinking PND 56: PND 60-66: LTP, Neuronal Morphology
Control (tap water), M, n water
= 7 11 |jg/L (1.1 |jg/dL) for Control
100 ppm, M, n = 9 133 |jg/L (13.3 |jg/dL) for 100
ppm
Li et al. (2016b)
Mouse (Kunming) GD0toPND21
Control (deionized
water), M/F, n = 10
0.1% (1000 ppm), M/F, n
= 10
0.2% (2000 ppm), M/F, n
= 10
0.5% (5000 ppm), M/F, n
= 10
Oral,
lactation
In utero
PND 21:
8.27 |jg/L (0.827 pg/dL) for
Control
41.05 ug/L (4.11 ug/dL) for
0.1% solution
82.93 ug/L (8.29 ug/dL) for
0.2% solution
105.33 |jg/L (10.53 pg/dL) for
0.5% solution
PND 21: Histopathology
Sobolewski et al. (2018)
Mouse (C57BL/6) GD -60 to PND 21
Oral,
PND 6-7:
PND 60: Epigenetics
Control (deionized
lactation
water), M/F, n = 6-12
In utero
0.37 pg/dL for Control
100 ppm, M/F, n = 6-12
10.2 pg/dL for 100 ppm
3-240
-------
Study Species (Stock/Strain), TimingoJ Exposurc BLL as Reported (Mg/dL) Endpoints Examined
Barkur and Bairv (2016)
Rat (Wistar)
GD -30 to PND21
Oral,
PND 22:
PND 30: Histopathology
Control (tap water with
lactation
acetic acid), M/F, n = 8
In utero
0.5 |jg/dL for Control
0.2% solution,
9.4 |jg/dL for 0.2% solution,
pregestation only (PG),
pregestation only (PG)
M/F, n = 8
16.6 |jg/dL for 0.2% solution,
0.2% solution, gestation
gestation only
only, M/F, n = 8
30.1 |jg/dL for 0.2% solution,
0.2% solution, lactation,
lactation
M/F, n = 8
33.4 |jg/dL for 0.2% solution,
0.2% solution, gestation
gestation and lactation
and lactation, M/F, n = 8
Shvachiv et al. (2018)
Rat (Wistar)
Intermittent
Oral, drinking
PND 196:
PND 189: Brain Histopathology
Control (tap water), M/F,
Exposure: GD 7 to
water
n = 8
PND 84, PND 140
Oral,
<0.1 |jg/dL for Control
to PND 196
lactation
0.2% (p/v) solution
In utero
18.8 |jg/dL for 0.2%
(distilled water), M/F, n =
Continuous
(Intermittent)
9 - Intermittent exposure
Exposure: GD 7 to
PND 196
24.4 |jg/dL for 0.2%
0.2% (p/v) solution, M/F,
(Continuous)
n = 9 - Continuous
exposure
Stansfield et al. (2015)
Rat (Long-Evans)
GD Oto PND 50
Oral, diet
PND 50:
PND 50: Neurotransmitter Analysis,
Control (chow), M/F, n =
Oral,
Brain Histopathology
4-7
lactation
0.6 |jg/dL for Control
In utero
1500 ppm, M/F, n = 4-7
22.2 |jg/dL for 1500 ppm
Listos et al. (2013)
Rat (Wistar)
GD Oto PND 28
Oral, drinking
PND 60:
PND 60: Neurotransmitter
Control (tap water), M/F,
water
n = 6-11
Oral,
0.93 |jg/dL for Control
lactation
0.1% solution, M/F, n =
In utero
20.45 |jg/dL for 0.1% solution
6-11
3-241
-------
Zhao etal. (2018)
Rat (Sprague Dawley) GD -14 to PND 10 Oral,
Control (tap water), M, n lactation
= 8 In utero
0.005% solution, M, n =
8
0.01% solution, M, n = 8
0.02% solution, M, n = 8
3-242
PND 0:
1.9 |jg/dL for Control
17.9 |jg/dL for 0.005% solution
23.2 |jg/dL for 0.01% solution
48.8 |jg/dL for 0.02% solution
PND 3:
I.9 |jg/dL for Control
6.7 |jg/dL for 0.005% solution
II.5 |jg/dL for 0.01% solution
23.1 |jg/dL for 0.02% solution
PND 7:
1.3 |jg/dL for Control
8.1 |jg/dL for 0.005% solution
12.3 |jg/dL for 0.01 % solution
18.7 |jg/dL for 0.02% solution
PND 10:
1.2 |jg/dL for Control
5.6 |jg/dL for 0.005% solution
7.0 |jg/dL for 0.01% solution
12.3 |jg/dL for 0.02% solution
PND 14:
0.7 |jg/dL for Control
PND 30: Electrophysiology,
Histopathology
-------
Study Species (Stock/Strain), TimingoJ Exposurc BLL as Reported (Mg/dL) Endpoints Examined
4.0 |jg/dL for 0.005% solution
5.5 |jg/dL for 0.01% solution
8.9 |jg/dL for 0.02% solution
PND 21:
1.1 |jg/dL for Control
2.5 |jg/dL for 0.005% solution
2.5 |jg/dL for 0.01% solution
2.98 |jg/dL for 0.02% solution
PND 30:
1.5 |jg/dL for Control
1.0 |jg/dL for 0.005% solution
1.5 |jg/dL for 0.01% solution
1.5 |jg/dL for 0.02% solution
PND 28 - Females: PND 28: Histopathology
0.02 |jg/dL for Control
3.03 |jg/dL for 30 ppm
12.79 |jg/dL for 330 ppm
PND 28 - Males:
0.03 |jg/dL for Control
3.68 |jg/dL for 30 ppm
15.42 |jg/dL for 330 ppm
Dominquez et al. (2019) Mouse (C57BL/6) PND 0 to PND 28 Oral,
Control (tap water), M/F, lactation
n = 10 (7/3)
30 ppm, M/F, n = 10
(6/4)
330 ppm, M/F, n = 10
(4/6)
3-243
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Du etal. (2015)
Rat (Sprague Dawley) PND 0 to PND 90
Control (distilled water),
M/F, n = 8
250 ppm, M/F, n = 8
Oral, drinking PND 30:
water
Oral, 13.9 pg/L (1.4 pg/dL) for
lactation Control
205.6 pg/L (20.6 pg/dL) for 250
ppm
PND 60:
15.0 pg/L (1.5 pg/dL) for
Control
321.9 pg/L (32.2 pg/dL) for 250
ppm
PND 90:
11.8 pg/L (1.2 pg/dL) for
Control
379.2 pg/L (37.9 pg/dL) for 250
ppm
PND 30, 60, 90: Histopathology
Mansouri et al. (2013)
Rat (Wistar) PND 55 to PND
Control (tap water or 181
water + NaAc), M/F, n =
16 (8/8)
50 ppm, M/F, n = 16
(8/8)
Oral, drinking PND 178-181 - Females:
water
NR for Control
10.6 pg/dL for 50 ppm
PND 178-181 - Males:
NR for Control
18.9 pg/dL for 50 ppm
PND 161-179: Neurotransmitter
3-244
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Zhou etal. (2018)
Rat (Sprague Dawley) PND 24 to PND 52 Oral, drinking PND 52:
Control (distilled water), water
M, n = 10
0.5% solution, M, n = 10
1.0% solution, M, n = 10
2.0% solution, M, n = 10
13.3 |jg/L (1.3 pg/dL) for
Control
148.9 |jg/L (14.9 pg/dL) for
0.5% solution
231.3 pg/L (23.1 pg/dL) for
1.0% solution
PND 24, 31, 38, 45, 52: Brain Weight,
Brain Histopathology
293.4 ug/L (29.3 pg/dL) for
2.0% solution
Dumkova et al. (2017) Mouse (ICR) NR(24g)-6wk Inhalation After 6 wk treatment: After 6 wk treatment: Histopathology
Control, F, n = 10 continuous
exposure 11 ng/g (1.16 pg/dL) for
106/cm3Pb0 Control
nanoparticles, F, n = 10
132 ng/g (13.99 pg/dL) for
106/cm3
3-245
-------
Study
Species (Stock/Strain), Timing of Exposure
n, Sex Exposure Details
BLL as Reported (pg/dL)
Endpoints Examined
Xiao etal. (2014)
Rat (Wistar) Pre-weaning: GD
Control (tap water), M/F, -21 to PND 21
n = 10 (5/5)
Postweaning: PND
Pre-weaning: 2 mM 21 to PND 84
solution, M/F, n = 10
(5/5)
Postweaning: 2 mM
solution, M/F, n = 10
(5/5)
Oral, drinking PND 21 - Pre-weaning:
water
10.09 |jg/L(1 pg/dL) for
Control
Oral,
lactation
In utero
PND 84 and PND 91: Histopathology
103.8 |jg/L (10.4 pg/dL) for 2
mM solution
PND 21 - Postweaning:
Not Reported
PND 91 - Pre-weaning:
10.32 pg/L (1 pg/dL) for
Control
39.27 pg/L (3.9 pg/dL) for 2
mM solution
PND 91 - Postweaning:
10.32 pg/L (1 pg/dL) for
Control
105.45 pg/L (10.5 pg/dL) for 2
mM solution
Sobin etal. (2013)
Mouse (C57BL/6) PND 1 to PND 28
Control (tap water), M/F,
n = 30
30 ppm, M/F, n = 30
230 ppm, M/F, n = 30
330 ppm, M/F, n = 30
Oral,
lactation
PND 28:
0.22 pg/dL for Control
4.12 pg/dL for 30 ppm
10.31 pg/dL for 230 ppm
13.84 pg/dL for 330 ppm
PND 28: Histopathology
3-246
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Sobolewski et al. (2020) Mouse (C57BL/6)
F0:
Control (deionized
water), F, n = 10
100 ppm, F, n = 10
F1:
see Figure 1, n = 12
F1: GD -60 to PND Oral,
23-27 lactation
In utero
F1 PND 6-7:
0 pg/dL for Control
12.5 pg/dL for 100 ppm (F0
dosing)
F3 PND 6-7:
0 ng/dL for Control
0 pg/dL for 100 ppm (F0
dosing)
PND 60-120 (variable by endpoint):
Neurotransmitter, Epigenetics
F2:
see Figure 1, n = 12
F3:
see Figure 1, n = 8-10
Ouvanq et al. (2019)
Rat (Sprague Dawley) GD0toPND679 Oral, drinking wk97:
Control (tap water), M/F,
n = 6-10
PND 679: Histopathology
0.05/0.01% solution,
M/F, n = 6-10
water
Oral,
lactation
In utero
0 mg/L (0 pg/dL) for Control
0.216 mg/L (21.6 pg/dL) for
0.05/0.01% solution
Saleh et al. (2019)
Rat (Sprague Dawley) NR (190-220g) -
Control (deionized 20 d of treatment
water), F, n = 8
Oral, drinking After 20 d treatment:
water
5.4 pg/dL for Control
After 20 d treatment: Brain Weight
Histopathology
160 ppm, F, n = 8
23.8 pg/dL for 160 ppm
3-247
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Singh etal. (2019)
Rat (Wistar)
Control (distilled water),
M, n = 5
2.5 mg/kg, M, n = 5
3 mo to 6 mo
Oral, gavage
6 mo:
5.76 pg/dL for Control
28.4 pg/dL for 2.5 mg/kg
6 mo: Brain Weight Brain
Histopathology
Xiao et al. (2020)
Rat (Sprague Dawley)
Control (tap water), F, n
= 10
125 ppm, F, n = 10
GD -7 to PND 68 Oral, drinking PND 68:
water
Oral,
lactation
In utero
PND 22, 68: Histopathology
24.23 ng/mL (2.4 pg/dL) for
Control
205 ng/mL (20.5 pg/dL) for 125
ppm
Sun et al. (2014)
Rat (Sprague Dawley)
Control (tap water), NR,
n = 20
580 ppm, NR, n = 20
NR (230-260 g) — 3 Oral, drinking
mo of treatment water
After 3 mo treatment: After 3 mo treatment: Histopathology
3.0 pg/L (0.3 pg/dL) for Control
56.8 pg/L (5.7 pg/dL) for 580
ppm
Su etal. (2016)
Rat (Sprague Dawley)
Control (deionized water
with 0.9% saline), M, n =
4
200 ppm, M, n = 4
PND 20 to PND 76 Oral, gavage PND 76:
PND 76: Histopathology
7.99 pg/L (0.8 pg/dL) for
Control
84.17 pg/L (8.4 pg/dL) for 200
ppm
Song etal. (2014)
Rat (Sprague Dawley)
Control (tap water), M, n
= 9
100 pg/mL, M, n = 9
200 pg/mL, M, n = 9
300 pg/mL, M, n = 9
PND 20-22 to PND Oral, drinking
76-78 water
PND 76-78:
0.73 pg/dL for Control
4.7 pg/dL for 100 pg/mL
10.1 pg/dL for 200 pg/mL
12.3 pg/dL for 300 pg/mL
PND 76-78: Histopathology
3-248
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Zhou et al. (2020a)
Rat (Sprague Dawley) GD 1 to PND 364
Control (distilled water),
M, n = 5-11
0.5 g/L solution, M, n =
5-11
2.0 g/L solution, M, n =
5-
Oral, drinking PND 21:
water
0 mg/L (0 pg/dL) for Control
0.1 mg/L (10 pg/dL) for 0.5 g/L
solution
0.36 mg/L (36 pg/dL) for 2.0
g/L solution
PND 364:
0 mg/L (0 pg/dL) for Control
0.15 mg/L (15 pg/dL) for 0.5
g/L solution
0.51 mg/L (51 pg/dL) for 2.0
g/L solution
PND 21, 364: Histopathology,
Electrophysiology
Liu et al. (2019)
Rat (Sprague Dawley) PND 1 to PND 21
Control (tap water), F, n
= 12
58 mg/L, F, n = 11
Oral,
lactation
PND 9:
0 pg/dL for Control
7.9 pg/dL for 58 mg/L
PND 21:
0 pg/dL for Control,
8.2 pg/dL for 58 mg/L
PND 40:
0 pg/dL for Control
0 pg/dL for 58 mg/L
PND 93: Histopathology
3-249
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Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Nan et al. (2016)
Mouse (C57BL/6)
Control (sterile water),
NR, n = 30
9.6 mmol/L, NR, n = 30
PND 21 to PND 56
Oral, drinking
water
PND 7:
0 |jg/L (0 pg/dL) for Control
106.3 |jg/L (10.6 pg/dL) for 9.6
mmol/L
PND 14:
0 |jg/L (0 pg/dL) for Control
293.2 |jg/L (29.3 pg/dL) for 9.6
mmol/L
PND 35:
0 |jg/L (0 pg/dL) for Control
959.6 |jg/L (96 pg/dL) for 9.6
mmol/L
PND 56: Histopathology
Singh et al. (2017)
Rat (Wistar) NR (160-200 g) -
Control (distilled water), 14 days of
M, n = 3-6 treatment
7.5 mg/kg, M, n = 3-6
Oral, gavage 12 hr after last treatment:
5.54 pg/dL for Control
30.28 pg/dL for 7.5 mg/kg
12 hr after last treatment: Brain Weight,
Histopathology
Biioor et al. (2012)
Rat (Wistar)
Control (deionized
water), M/F, n = 10
50 ppm, M/F, n = 10
GD Oto PND 45
Oral, drinking
water
Oral,
lactation
In utero
PND 45:
4.06 pg/dL for Control
10.65 pg/dL for 50 ppm
PND 45: Neurotransmitter
3-250
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Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Wanq et al. (2021a)
Rat (Sprague Dawley)
GD Oto PND21
Oral,
PND 21:
PND 21: Histopathology
Control (deionized
lactation
water), M, n = 3
In utero
23.1 |jg/L (2.31 pg/dL) for
Control
0.05% solution, M, n = 3
248 pg/L (24.8 pg/dL) for
0.1% solution, M, n = 3
0.05% solution
302 pg/L (30.2 pg/dL) for 0.1%
solution
361 pg/L (36.1 pg/dL) for 0.2%
solution
Liu et al. (2022c)
Rat (Sprague Dawley)
PND 35 to PND
Oral, drinking
PND 119:
PND 119: Histopathology
Control (tap water), M, n
119
water
= 10
10.9 pg/L (1.09 pg/dL) for
Control
0.2% solution, M, n = 10
176 pg/L (17.6 pg/dL) for 0.2%
solution
Hsu et al. (2021)
Rat (Sprague Dawley)
PND 42 to PND 77
Oral, drinking
PND 84:
PND 78 to PND 84: Electrophysiology
Control (deionized
water
water), M, n = 6
0.9 pg/L (0.09 pg/dL) for
Control
250 ppm, M, n = 6
15.3 pg/L (1.53 pg/dL) for 250
ppm
Sadeqhi et al. (2021)
Rat (Wistar)
GD Oto PND 50
Oral, drinking
PND 50:
PND 50: Histopathology
Control (untreated), M, n
water
= 5
Oral,
0.58 pg/dL for Control
lactation
1500 ppm, M, n = 5
In utero
3.4 pg/dL for 1500 ppm
Viaueras-Villasenor et al.
Rat (Wistar)
GD 0 to PND 21
Oral,
PND 110:
PND 90 to PND 110: Histopathology
(2021)
Control (tap water), M, n
lactation
= 20
In utero
2.04 pg/dL for Control
320 ppm, M, n = 20
26.3 pg/dL for 320 ppm
3-251
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Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
^etaMs'0 BLL as RePorted (MQ/dL)
Endpoints Examined
Long et al. (2022)
Rat (Sprague Dawley) 6 wk to 18 wk
Control (untreated), M, n
= 12
200 mg/L solution, M, n
= 12
Oral, drinking 18wk:
water
2.14 [jg/L (0.214 pg/dL) for
Control
32.48 |jg/L (3.25 pg/dL) for 200
mg/L solution
NR: Histopathology, Neurotransmitter
Abazvan et al. (2014)
Mouse (CAMKII-tTA;
heterozygous or
homozygous for
mDISCI)
Control (het), M/F, n =
5-10
Control (mutant), M/F, n
= 5-10
1500 ppm (het), M/F, n =
5-10
1500 ppm (mutant), M/F,
n = 5-10
GD Oto PND 180
Oral, diet 6 mo - Females:
Oral,
lactation, in 0.6 pg/dL for Control (het)
utero
0.8 pg/dL for Control (mutant)
34.9 pg/dL for 1500 ppm (het)
33.3 pg/dL for ppm (mutant)
6 mo - Males:
1.1 pg/dL for Control (het)
1.1 pg/dL for Control (mutant)
26.1 pg/dL for 1500 ppm (het)
25.0 pg/dL for 1500 ppm
(mutant)
PND 180: Brain Volume, Brain MRI,
Morphometric Measurements in several
regions
3-252
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Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
^etaMs'0 BLL as RePorted (MQ/dL)
Endpoints Examined
Zhu et al. (2013)
Rat (Sprague Dawley) GD 0 to PND 490
Control (untreated), M, n
= 11-13
510 mg/L, M, n = 11-13
Oral, drinking PND 21
0 |jg/L for control
water
Oral,
lactation, in
utero
0.27 mg/L (27 pg/dL) for 510
mg/L
PND 287
0 |jg/L for control
0.24 mg/L (24 pg/dL) for 510
mg/L
PND 490
0 |jg/L for control
PND 21 to PND 490: Histopathology
0.25 mg/L (25 pg/dL) for 510
mg/L
Nam et al. (2018a)
Rat (Sprague Dawley) GD 0 to PND 21
Control (distilled water),
M/F, n = 12
0.2% solution M/F, n =
12
Oral, drinking PND 21
1.28 pg/dL for control
PND 21: Histopathology
water
Oral
lactation, in
utero
12.67 pg/dL for 0.2% solution
Gassowska et al. (2016a) Rat (Wistar)
Control (drinking water),
M/F, n = 4-8
0.1% solution M/F, n =
4-8
GD 0 to PND 28 Oral, drinking PND 28
0.93 pg/dL for control
PND 28: Histopathology
water
Oral
lactation, in
utero
6.86 pg/dL for 0.1%
3-253
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Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Gassowska et al. (2016b) Rat (Wistar) GD0toPND28
Control (drinking water),
M/F, n = 4-8
0.1% solution, M/F, n =
4-8
Oral, drinking PND 28
0.93 pg/dL for control
PND 28: Histopathology
water
Oral
lactation, in
utero
6.86 pg/dL for 0.1%solution
Sepehri and Ganii (2016) Rat (Wistar)
Control, M, n = 8
0.05% solution, M, n = i
GD 5 to PND 25 Oral, drinking PND 25
water 0.78 pg/dL for control
Oral
lactation, in 28.3 pg/dL for 0.05%
utero
PND 25: Histopathology
Zhu et al. (2019a)
Rat (Sprague Dawley) GD -10 to 12 mo
Control (distilled water),
M, n = 10
0.5 g/L, M, n = 10
Oral, drinking 12 mo
water 0 pg/dL for control
Oral
lactation, in 0.27 mg/L (27 pg/dL) for 0.5
utero g/L
12 mo: Electrophysiology
Zhanq et al. (2015b)
Rat (Long-Evans)
GD-10 to PND 50 Oral, diet
PND 50
PND 50: Histopathology,
Control (0 ppm), M/F, n
Oral
0.8 pg/dL for control
Electrophysiology
= 10
lactation, in
utero
21.1 pg/dL for 1500 ppm
1500 ppm, M/F, n = 10
Wang, 2021,
Rat (Sprague Dawley)
GD-28 to PND 21 Oral,
PND 21:
10296633@@author-
Control (deionized
lactation
year
water), M/F, n = 12
In utero
23.9 pg/L (2.39 pg/dL) for
Control
0.05% solution, M/F, n =
10
206 pg/L (20.6 pg/dL) for
0.05% solution
3-254
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL) Endpoints Examined
Wu et al. (2020a)
Mouse (C57BL/6J)
GD -7 to 7 mo
Oral, diet
PND 207-210 7 mo: Brain Weight, Histopathology
Control (ultra-pure
water), F, n = 8
Oral
lactation, in
utero
Mouse (C57BL/6J)
19.71 |jg/L (1.97 pg/dL) for
control
200 mg/L, F, n = 8
84.53 pg/L (8.45 pg/dL) for 200
Mouse (APP/PS1)
mg/L
Control (ultra-pure
water),F, n = 8
Mouse (APP/PS1)
19.96 pg/L(1.99 pg/dL) for
control
200 mg/L, F, n = 8
205.49 pg/L(20.54 pg/dL) for
200 mg/L
Mani et al. (2020) Rat (Wistar) 8 mo to 9 mo Oral, gavage NR NR: Histopathology, Brain Weight
Control (distiller water), 2.3 pg/dL for Control
M, n = NR
8.5 |jg/dL for 10 mg/kg
10 mg/kg, M, n = NR
16.4 |jg/dL for 50 mg/kg
50 mg/kg, M, n = NR
16.3 |jg/d for 100 mg/kg
100 mg/kg, M, n = NR
3-255
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Study
Species (Stock/Strain), Timing of
n, Sex Exposure
^etaMs'0 BLL as RePorted (MQ/dL)
Endpoints Examined
Shvachiv et al. (2020)
Rat (Wistar) GD 7 to 3 mo
Control (tap water), M/F,
GD 7 to 5 mo
n = 12
0.2% solution (3 mo)
M/F, n = 12
0.2% solution (5 mo)
M/F, n = 12
0.2% solution (7 mo)
M/F, n = 12
GD 7 to 7 mo
Oral, drinking 3 mo:
water
Oral,
lactation
In utero
<1 for Control
24.0 |jg/dL for 0.2% solution
5 mo:
<1 for Control
24.8 |jg/dL for 0.2% solution
PND 189: Histopathology
7 mo:
<1 for Control
26.9 |jg/dL for 0.2% solution
3-256
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Zhao etal. (2021)
Rat (Sprague Dawley) GD-14toPND10 Oral, drinking PND 10
water
Oral,
lactation
In utero
Control, M, n = 6
109 ppm, M, n = 6
0.6 pg/dL for control
11.4 pg/dL for 109 ppm
PND 21
0.85 pg/dL for control
3.5 pg/dL for 109 ppm
PND 30
0.98 pg/dL for control
PND 30: Histopathology,
Electrophysiology
1.8 pg/dL for 109 ppm
APP = amyloid precursor protein; BLL = blood lead level; Dl = deionized; F = female; F0 = gestating female; GD = gestational day; LTP = long-term potentiation; M = male; ME =
maternal exposure; MRI = magnetic resonance imaging; mo = month(s); NaAc = sodium acetate; NR = not reported; Pb = lead; PG = pregestation; PND = postnatal day; PW =
postweaning; wk = week(s); yr = year(s).
3-257
-------
Table 3-2E Epidemiologic studies of Pb exposure and full-scale intelligence quotient
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Lanphear et al. (2005)
Lanphear et al. (2019)
International pooled
analysis: Prospective
cohorts from Boston,
Cincinnati, Cleveland,
Mexico City, Port Pirie,
Rochester, and
Yugoslavia.
Followed from birth
(1979-1995) up to age 10
yr
n = 1,333 children
Blood
Median (5th-95th)
Early childhood (6-24
mo):
12.7 (3.5-34.5)
Peak: 18 (6.0-47.0)
Lifetime avg (through
outcome measurement at
4.8-10 yr):
11.9 (3.6-34.5)
Concurrent: 9.7 (2.5-
33.2)
FSIQ: WISC-III, WISC-R, HOME score, birth weight, Early Childhood:
WPPSI, WISC-S
maternal IQ and
(depending on the cohort) education. Also
Ages 4.8-10 yr
considered potential
confounding by child sex,
birth order, marital status,
maternal age, prenatal
smoking status and
alcohol use.
-0.137 (-0.209, -0.064)
Lifetime avg:
-0.206 (-0.285, -0.126)
Concurrent:
-0.187 (-0.26, -0.114)
Peak:
-0.126 (-0.182, -0.071)
Lanphear et al. (2005)
Lanphear et al. (2019).
subset of with peak BLLs
<7.5 |jg/dL
n = 103 children
Same
Concurrent
Mean: 3.2
Same
Same
-2.53 (-4.48, -0.58)
3-258
-------
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
al
Crump et al. (2013) N = 1355 children (note Blood
that the reanalysis
„ , , x included prenatal BLLs in
Reanalysis ofLanphearet |jfetj d , .
12019, 2005) a different strategy P '
including covariates that
allowed a larger number
of observations to be
included in the analysis)
FSIQ: WISC-III, WISC-R,
WPPSI, WISC-S
Same but defined to be
cohort specific
BLL distribution reported
by cohort not for the
pooled dataset as a
whole.
Concurrent: (3= -3.315
(-4.546, -2.084)
Peak
(3= -2.484 (-3.825,
-1.142)
Early childhood
(3= -2.459 (-3.817,
-1.102)
Lifetime avg
(3= -3.246 (-4.659,
-1.833)
24-month
(3= -1.955 (-3.193,
-0.717)
Note: the estimates are
not standardized)
Van Landinaham et al. NR
(2020)
Reanalysis of Lanphear et
al. (2005) and Lanphear
et al. (2019)
HOME x In (BLL+1):
0.0437
NR FSIQ: WISC-III, WISC-R, Defined highly likely B=-4.945
WPPSI, WISC-S confounders: HOME Interaction terms'
score, maternal education ,, x
and maternal IQ and Maternal IQ x ""(BLL+1)
included interaction terms -0.0003
between BLL and each of Mother's education x
these covariates inmi i+-iv .n rrai
Note: the estimates are
not standardized)
Canfield et al. (2003a)
n = 101
Blood
FSIQ
Child sex, Fe status, birth -1.8 (-3.0, -0.60)
Rochester, NY
Children recruited from
Stanford-Binet
weight, maternal race,
Prospective cohort
dust control study
Concurrent, children with
peak <10
Age 5 yr
education, IQ, income,
and prenatal smoking
Born 1994-1995 followed
status, HOME score.
from age 6 mo to 5 yr
Mean: 3.3
3-259
-------
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Bellinger and Needleman n = 48 children
(2003) Recruited at birth
Boston area, MA.
Prospective
Followed from birth
(1979-1981) to age 10 yr.
Blood
Early childhood (age 2 yr)
Mean (SD)
Peak <10: 3.8 (range: 1-
9.3)
Detection limit NR
FSIQ: WISC-R
Age 10 yr
HOME score (age 10 and
5), child stress events,
race, maternal IQ, age,
marital status, SES, sex,
birth order, # residence
changes before age 5 yr.
Also considered potential
confounding by family
stress, maternal age,
psychiatric factors, child
serum ferritin levels
-1.6 (-2.9, -0.2)
Surkan et al. (2007)
Boston, MA and
Farmington, ME
Cross-sectional
Sep 1997-Mar 2005
n = 389
Children recruited from
trial of amalgam dental
fillings.
Blood
Concurrent
Groupl: 1-2
Group 2: 3-4
Group 3: 5-10
Mean (SD):
2.2(1.6)
WISC-III
Age 6-10 yr
Caregiver IQ, child age,
SES, race, birth weight.
Also considered site, sex,
birth order, caregiver
education and marital
status, parenting stress,
and maternal utilization of
prenatal and annual
health care (not parental
caregiving quality.)
1.0 (reference)
-0.12 (-3.3, 3.1)
-6.0 (-11, -1.4)
Chiodo et al. (2007)
Detroit, Ml area
Cross-sectional
495 children (born 1989-
1991) age 7 yr,
Blood
Concurrent
Mean (SD): 5.0 (3.0)
WPPSI
Age 7 yr
Maternal
psychopathology, IQ,
prenatal smoking,
prenatal marijuana, SES,
HOME score, caretaker
education and marital
status, # children in
home, child sex. Also
considered child age,
maternal age, custody,
cocaine use, prenatal
alcohol use.
-0.19 (-0.30, -0.08)
Note: standardized
regression coefficient.
95% CI estimated using
the reported p-value of
0.01
3-260
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Kim et al. (2009) 261 children
Seoul, Seongnam, Ulsan,
and Yeoncheon, South Schoo, reCruitment
Korea
Cross-sectional
Children born 1996-1999
Blood
Concurrent
Mean (SD): 1.7 (0.80)
Age: 8-11 yr
KEDI-WISC
Ages 8-11 yr
Blood Mn <1.4 |jg/dL
-2.4 (-6.0, 1.1)
Maternal age, education
and prenatal smoking
status, paternal
education, yearly income, Blood Mn >1-4 M9/c|L
smoking exposure status ~3.2 (-6.1, -0.24)
after birth, child age, sex,
and birth weight (not
parental caregiving quality
or IQ)
tBraun et al. (2018)
Cincinnati, OH
United States
Mar 2003-Jan 2006
Followed for 8 yr
Cohort
HOME study
n: 355 (Intervention
group: 174, Control
group: 181)
Clinical trial of pregnant
women, mean gestation
of 16 wk and residence in
a house built in or before
1978
Intervention to reduce Pb
exposure
Blood
Maternal and child blood;
ICP-MS
Dust Pb loadings floor,
interior windowsill and
window at 20 wk
gestation, child age 1 and
2 yr; GFAAS.
Age at Measurement:
16, 26 wk of gestation,
delivery (maternal);
1,2,3,4,5, 8 yr (child)
Baseline GM (Intervention
and control groups):
maternal: 0.7 and 0.7
|jg/dL, floor dust Pb: 1.5
and 1.9 pg/sq ft,
windowsill dust Pb: 28
and 33 pg/sq ft, window
trough dust Pb: 574 and
510 |jg/sq ft.
WPPSI
Age at outcome:
5-8 yr
NA
Mean FSIQ score
difference15: 0.5 (-3.3,
24.2), comparing the
treatment to the injury
prevention control group.
3-261
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
t Taylor et al. (2017)
UK
Cohort
April 1, 1991-Dec 31,
1992 (followed until age
4-8 yr)
ALSPAC
n: 4285
Mother-infant pairs.
Blood FSIQ, VIQ, PIQ (WPPSI
or WISC-111).
Maternal and child venous
blood; ICP-MS Age at outcome:
4-8 yr
Age at Measurement:
Prenatal (mean
gestational age 11 wk)
and postnatal (30 mo)
Prenatal: 3.67 |jg/dL;
Child BLL: 4.22 pg/dL.
Family adversity index,
housing tenure,
household crowding,
smoking in the first
trimester, alcohol
consumption in the first
trimester, maternal age at
index birth, parity,
maternal education,
length of time the mother
lived in Avon, child sex,
child age at testing,
weighted life events
score, and hemoglobin
level.
WISC-Boysb: -0.29
(-1.02, 0.44)
WISC-Girlsb: 0.73 (0.13,
1.33)
WPPSI-Girlsb: -0.65
(-2.065, 0.765)
WPSSI-Boysb: -0.54
(-2.015, 0.935)
tTatsuta et al. (2020)
Tohoku district, coastal
area Japan
2002-2006 (enrollment)
through 2015-2018 (12 yr
followup)
Cohort
TSCD coastal cohorts
n: 289 mother-child pairs
(singleton births); 148
boys and 141 girls
Blood
Cord and child venous
blood; ICP-MS
Age at Measurement:
Delivery (cord), 12 yr
(child)
Median: Cord = 0.8 pg/dL,
12-yr = 0.7 pg/dL
95th: Cord: 1.4 pg/dL, 12-
yr: 1.1 pg/dL
FSIQ (Japanese version
WISC-IV), age equivalent
ranking and scores for
verbal comprehension,
perceptual reasoning,
working memory, and
processing speed
composites; BNT (cues
and no cues)
Age at outcome:
12 yr
Birth weight, drinking or
smoking during
pregnancy, the Raven's
score (parent assessment
for child at 18 mo of age),
passive smoking status at
12 yr old, family income,
WISC/BNT tester, and
cord blood total Hg
Cord-Boys: p=—3.683 (-
Cord-Girls: (3 = 10.714,
3.349)
Child-Boys: (3 = 1.463
(-2.905, 5.831)
Child-Boys: (3 = -9.88
(-18.977, -0.782)
Child-Girls: -4.406
(-15.94, 7.129)
3-262
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tDesrochers-Couture et
al. (2018)
10 study sites
Canada
2008-2011
Followed 3-4 yr from birth
Cohort
MIREC Study Blood
n: 609
Maternal, cord, and
Birth cohort: Mother-infant postnatal child (venous)
pairs recruited during 1st ICP-MS
trimester Age at Measurement:
Maternal (6-13 wk, 32-34
wk), birth (cord); and 3-4
yr (postnatal child)
GM: 1st trimester: 0.62
|jg/dL; 3rd trimester: 0.59
|jg/dL; cord blood: 0.76
|jg/dL; child blood: 0.70
pg/dL
Max: 1st trimester: 4.14
|jg/dL; 3rd trimester: 3.93
|jg/dL; cord blood: 3.52
|jg/dL; child blood: 5.49
Mg/dL.
FSIQ, VIQ, PIQ, General
Language composite
WPPSI-3rd Edition, short
version. The age-
standardized WPPSI-III
Canadian norms
were used to calculate the
scores.
Age at outcome:
Between 2 yr 6 mo and 3
yr 11 mo
Cord blood model: child
age, child sex, maternal
education, evaluation site
and cord blood Hg (log—2
scale).
Child blood model: child
age, child sex, evaluation
site, marital status,
familial income, HOME
total score, Parenting
Stress Index, and cord
blood Pb (log—2 scale).
Cord: (3=-0.123 (-0.251,
0.005)
Child: (3 = 0.027 (-0.135,
0.188)
Cord-Boys: p=—5.686
(-9.968, -1.405)
Cord-Girls: (3 = 0.287
(-3.787, 4.361)
tZhou et al. (2020b)
Jiangsu Province
China
June 2009-Jan 2010 to
June 2016-July 2017
Cohort
Sheyang Mini Birth
Cohort Study
n: 296
Blood, Urine
FSIQ, VIQ, PIQ (Chinese Sex, maternal age,
Cord blood tested for Mn,
Cd and Pb using GFAAS.
Birth cohort- mother-infant Postnatal urine samples
pairs from an agricultural urine also tested for the
region. elements.
Age at Measurement:
Cord; postnatal Urine NR
GM: cord blood: 15.88
|jg/L, urine: 1.43 |jg/L,
75th: cord blood: 21.83
|jg/L, urine: 2.27 |jg/L,
Max: cord blood: 1168.20
|jg/L, urine: 62.47 |jg/L,
version WISC-R).
Age at outcome:
6-7 yr (school-aged
children)
maternal education,
family annual income,
family inhabitation area,
and passive smoking;
multiple effects were also
assessed by entering all
other metals in the model.
Sex-stratified analysis
conducted.
Cord-Girls: 0.615 (-0.909,
2.138)
Cord-Boys: 0.835
(-1.164, 2.833)
Cord-All: 0.67 (-0.514,
1.854)
3-263
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Referencejmd Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tLiu etal. (2015)
Chongqing, China
March 4, 2003-June 19,
2003 (enrollment)
Followed 5 yr
Cohort
Birth cohort with mother-
infant pairs
n: 149
Mothers gave birth in 4
hospitals in Tongliang
county.
Blood
VIQ, PIQ, and FSIQ
(Shanghai version
WPPSI)
Maternal and cord serum
(Hg, Cd, Pb); Pb
determined with AAS.
Ratio of maternal to cord
serum levels (i.e.,
placental transport ratio of equivalent percentile
metal) estimated. ranks were estimated
Raw scores were
converted to composite
scores and age
Age at Measurement:
Delivery
Age at outcome:
5 yr
Maternal age, educational
level, vitamins, placental
transport ratios, maternal
exposure to ETS. Pb, Hg
and Cd were considered
in multivariable models
(covariates retained
based on evaluation of
VIFs).
Final predictive model did
not include associations
with Pb
Mean: 3.45 |jg/dL.
3-264
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tWana et al. (2022) n: 148
Blood
Cognitive Effects in
General linear models
"Table 3, Beta (95% CI):
children - FSIQ
adjusted for child sex,
Wujiang
Blood Pb measured via
maternal age at delivery,
VIQ, total population, cord
China
AAS.
Child health personnel
age of children, maternal
blood:
Birth cohort established
Age at Measurement:
conducted the WISC-CR
education level, paternal
from 2009-2010; follow-
Mean, SD (mo): 89.90,
for FSIQ, PIQ, and VIQ.
education level, monthly
Q2 vs. Q1: -0.296
up from 2016-2017.
3.77
Scores were age
household income, parity,
(-7.005, 6.413)
Cohort
converted and
inhabitation area, passive
Cord blood: GM = 28.26
standardized.
smoking.
Q3 vs. Q1: 4.468 (-2.840,
|jg/L, Median = 27.56
11.776)
|jg/L; Venous blood: GM =
Age at outcome:
22.99 |jg/L, Median =
Q4 vs. Q1: -1.275
23.80 |jg/L
(-8.231, 5.682)
75th: Cord blood: 38.42
|jg/L; Venous blood:
VIQ, boys, cord blood:
33.00 |jg/L
Max: Cord blood: 249.00
Q2 vs. Q1: -2.752
|jg/L; Venous blood:
(-12.594, 7.091)
71.40 |jg/L.
Q3 vs. Q1: 5.681 (-6.288,
17.649)
Q4 vs. Q1: 2.119 (-8.913,
13.150)
VIQ, girls, cord blood:
Q2 vs. Q1: 5.679 (-5.242,
16.600)
Q3 vs. Q1: 6.510 (-4.501,
17.521)
Q4 vs. Q1: -1.331
(-11.933, 9.271)
VIQ, total population,
venous blood:
Q2 vs. Q1: 4.942 (-3.957,
13.841)
Q3 vs. Q1: -0.536
(-9.560, 8.487)
3-265
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3-266
Q4 vs. Q1: -3.304
(-12.117, 5.509)
VIQ, boys, venous blood:
Q2 vs. Q1: 5.592 (-7.193,
18.378)
Q3 vs. Q1: 8.858 (-4.271,
21.988)
Q4 vs. Q1: 7.143 (-5.649,
19.935)
VIQ, girls, venous blood:
Q2 vs. Q1: 3.179
(-10.810, 17.169)
Q3 vs. Q1: -13.548
(-27.506, 0.411)
Q4 vs. Q1: -14.964
(-28.412, -1.517), p-
value = 0.036
PIQ, total population, cord
blood:
Q2 vs. Q1: -5.584
(-14.011, 2.842)
Q3 vs. Q1: -0.441
(-9.620, 8.738)
Q4 vs. Q1: -9.365
(-18.103, -0.628), p-
value = 0.038
PIQ, boys, cord blood:
Q2 vs. Q1: -13.080
(-24.907, -1.254), p-
-------
value = 0.035
Q3 vs. Q1: -9.686
(-24.067, 4.695)
Q4 vs. Q1: -7.592
(-20.848, 5.663)
PIQ, girls, cord blood:
Q2 vs. Q1: -0.002
(-14.192, 14.188)
Q3 vs. Q1: 4.023
(-10.284, 18.331)
Q4 vs. Q1: -13.293
(-27.069, 0.483)
PIQ, total population,
venous blood:
Q2 vs. Q1: -7.293
(-17.605, 3.020)
Q3 vs. Q1: -8.176
(-18.633, 2.281)
Q4 vs. Q1: -4.507
(-14.720, 5.706)
PIQ, boys, venous blood:
Q2 vs. Q1: -8.218
(-22.276, 5.841)
Q3 vs. Q1: -6.701
(-21.138, 7.737)
Q4 vs. Q1: -4.294
(-18.360, 9.772)
PIQ, girls, venous blood:
3-267
-------
3-268
Q2 vs. Q1: -10.417
(-29.104, 8.270)
Q3 vs. Q1: -16.397
(-35.043, 2.248)
Q4 vs. Q1: -6.994
(-24.957, 10.968)
FSIQ, total population,
cord blood:
Q2 vs. Q1: -3.369
(-10.711, 3,974)
Q3 vs. Q1: 2.396 (-5.603
10.394)
Q4 vs. Q1: -6.087
(-13.700, 1.527)
FSIQ, boys, cord blood:
Q2 vs. Q1: -8.599
(-19.015, 1.818)
Q3 vs. Q1: -1.434
(-14.100, 11.232)
Q4 vs. Q1: -2.552
(-14.227, 9.123)
FSIQ, girls, cord blood:
Q2 vs. Q1: 2.986 (-9.126
15.098)
Q3 vs. Q1: 5.743 (-6.469
17.955)
Q4 vs. Q1: -8.635
(-20.394, 3.123)
FSIQ, total population,
-------
Referencejmd Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
venous blood:
Q2 vs. Q1: -0.950
(-10.606, 8.706)
Q3 vs. Q1: -4.930
(-14.722, 4.861)
Q4 vs. Q1: -4.773
(-14.336, 4.790)
FSIQ, boys, venous
blood:
Q2 vs. Q1: -1.367
(-14.675, 11.941)
Q3 vs. Q1: 1.366
(-12.300, 15.033)
Q4 vs. Q1: 1.929
(-11.386, 15.243)
FSIQ, girls, venous blood:
Q2 vs. Q1: -3.771
(-20.071, 12.529)
Q3 vs. Q1: -17.326
(-33.590, -1.062), p-
value = 0.044
Q4 vs. Q1: -13.625
(-29.293, 2.0"3)"
3-269
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tlalesias et al. (2011)
Antofagasta,
Northern Chile
1998-2005
Cohort
Sepulveda study
n: 192
Children lived or attended
school in area
contaminated by Pb
mineral concentrate
stored in open sites at
railroad terminal (closed
in 1998).
Blood
Child's venous blood;
AAS.
Age at Measurement:
In 1998 at 0-7 yr old; in
2005 at 7-16 yr old
Mean and Median: blood
Pb 1998: 10.8 and 10
|jg/dL; blood Pb 2005: 3.5
and 3.2 |jg/dL
75th: blood Pb 1998: 14
|jg/dL; blood Pb 2005: 4.3
|jg/dL
Max: Blood Pb 1998: 33
|jg/dL; blood Pb 2005: 14
(jg/dL.
FSIQ, VIQ, PIQ (Chilean
version of WISC-R).
Age at outcome:
7-16 yr
Sex, birth weight, birth
order, # of siblings, milk
type during the first six
months of life, history of
anemia, SES (household
income, home ownership,
and school type: public or
private), parental
education, maternal
smoking during
pregnancy, maternal IQ,
children's stimulation at
home, HOME score.
Concurrent: -0.94 (-1.77,
-0.11)
Early childhood: -0.14
(-0.445, 0.165)
tRuebner et al. (2019)
46 centers
U.S.
Cohort
3 enrollment periods,
2005-2009, 2011-2014,
2016-2020
Followed up to 9 yr
CKiD Cohort study
n: 412
Children with mild to
moderate CKD
Blood
ICP-
Child venous blood
MS. The BLL
measurement closest to
the time of neurocognitive
testing was used for
analysis (concurrent).
Age at measurement:
NR; 2, 4, or 6 yr after
study entry
Median: 1.2 |jg/dL
75th: 1.8 |jg/dL
Max: 5.1 |jg/dL
FSIQ
Mullen Scales of Early
Learning (age 12-29 mo),
WPPSI (30 mo-5 yr), and
WAS I (6-18 yr).
The last available test
results were to evaluate
long-term effects. Mean
time between BLL and
neurocognitive testing
was 2.3 yr.
Age at outcome:
1-16 yr
Age, sex, race, poverty,
and maternal education
Concurrent: (3=-2.1
(-3.95, -0.25)
3-270
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tLee etal. (2021)
Seoul, Gyeonggi, and
Incheon provinces
South Korea
2008-2017 (Recruitment
2008-2010; follow-up
from 2012-2017)
Cohort
Environment and
Development of Children
n: 502
Blood
Cognitive Effects (FSIQ)
Intelligence quotients of
children assessed using
'KEDI-WISC.
Whole blood Pb from
mothers during their
second trimester of
pregnancy and children at
ages 4 yr and 6 yr were Age at outcome:
analyzed by atomic
absorption
spectrophotometry.
Age at Measurement:
Maternal mean age (SD)
= 31.3 (3.5) yr. Children at
4 yr and 6 yr.
Prenatal GM (SD) = 1.32
(1.32) |jg/dL; children at 4
yr = 1.43 (1.38) pg/dL;
children at 6 yr = 1.43
(1.35) pg/dL
75th: Prenatal = 1.56
pg/dL; children at 4 yr =
1.72 pg/dL; children at 6
yr = 1.70 pg/dL
95th: Prenatal = 2.11
pg/dL; children at 4 yr =
2.47 pg/dL; children at 6
yr = 2.28 pg/dL.
Multivariate models
adjusted by maternal
education level, exposure
to ETS during the
pregnancy, maternal age,
and maternal IQ.
Table-3 - Estimated
coefficients and 95% CI of
associations between
single metals and
children's IQ at 6-years
old (standardized)
Prenatal period: -1.202
(-4.87, 2.467)
At age 4: -1.829 (-4.664,
1.006)
At age 6: -2.614 (-5.623,
0.396)
tDantzer et al. (2020)
Greater Cincinnati, Ohio
Metro area
U.S.
Cross-sectional analysis
of data collected at age
12
CCAAPS
n: 344
Cohort recruited at birth
Oct 2001 —Jul 2003
Blood, nails
Postnatal child venous
blood, toenail; ICP-MS.
Mean blood Pb: 0.57
pg/dL; toenail Pb: 0.66
pg/g; information also
available by gender and
race
Age: 12 yr.
FSIQ (WISC-IV)
Age at outcome:
12 yr
Caregiver IQ, community
deprivation index, and
BMI. Sex considered as a
potential confound.
Concurrent (blood):
B=-10.871 (-16.893,
-4.848)
B=-1.70 (-4.27, -0.862)
3-271
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tMartin et al. (2021)
East Liverpool, Ohio
United States
2013-2014
Cross-Sectional
CARES
n: 66
BLLs from the children
were analyzed by ICP-
MS.
Age at Measurement:
Mean (SD) = 8.4 (0.9) yr
GM (SD) = 1.13 (1.96)
|jg/dL
Max: 6.64 |jg/dL.
Cognitive Effects
Cognitive performance
was assessed using the
WISC-IV.
Age at outcome:
Regression models were
adjusted for sex, income,
and In (serum cotinine).
"Table 3 - Interaction
effects between Ln Blood
Pb-Ln Hair Mn, 13 (95%
CI)
Blood Pb (per 1 In |jg/dL
difference)
FSIQ
At In hair Mn = 5 ng/g:
1.686 (-3.039, 6.412)
At In hair Mn = 6.25 ng/g:
-4.447 (-8.333, -0.561)
At In hair Mn = 7 ng/g:
-8.133 (-13.4, -2.867)
At In hair Mn = 7.5 ng/g:
-10.596 (-17.172, -4.02)
Perceptual Reasoning
At In hair Mn = 5 ng/g:
2.392 (-4.055, 8.839)
At In hair Mn = 6.25 ng/g:
-4.612 (-9.918, 0.694)
At In hair Mn = 7 ng/g:
-8.808 (-15.996, -1.62)
At In hair Mn = 7.5 ng/g:
-11.616 (-20.592,
-2.639)
Processing Speed
At In hair Mn = 5 ng/g:
-0.62 (-4.263, 3.024)
At In hair Mn = 6.25 ng/g:
-2.565 (-5.565, 0.435)
3-272
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Referencejmd Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
At In hair Mn = 7 ng/g:
-3.725 (-7.788, 0.337)
At In hair Mn = 7.5 ng/g:
-4.502 (-9.576, 0.573)
Verbal Comprehension
At In hair Mn = 5 ng/g:
1.208 (-3.451, 5.867)
At In hair Mn = 6.25 ng/g:
-4.141 (-7.976, -0.306)
At In hair Mn = 7 ng/g:
-7.349 (-12.549, -2.149)
At In hair Mn = 7.5 ng/g:
-9.49 (-15.976, -3.004)
Working Memory
At In hair Mn = 5 ng/g:
2.376 (-2.404, 7.157)
At In hair Mn = 6.25 ng/g:
-2.133 (-6.067, 1.8)
At In hair Mn = 7 ng/g:
-4.831 (-10.161, 0.498)
At In hair Mn = 7.5 ng/g:
-6.635 (-13.286, 0.0"6)"
3-273
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tHavnes et al. (2015)
Marietta or Cambridge,
Ohio, and surrounding
communities
U.S.
Oct 2008-M arch 2013
Cross-sectional
CARES
n: 404
Blood
Child venous blood: ICP-
MS
Age at Measurement:
Participants resided in
study area throughout
their life; not moving for at -jIq yr
least 1 yr
GM: 0.82 pg/dL
Max: NR.
FSIQ (WISC-IV), 4
domains of intellectual
functioning (reasoning,
processing speed,
working memory and
verbal comprehension)
Age at outcome:
7-9 yr
Sex, parent's IQ, parent Processing speed: (3 = -
education, parent 3.53 [ -6.95, -0.12)
confidence T-score, Mn or
Pb, community residence Association with FSIQ NR
(FSIQ models only); other
sets of variables added
depending on domain.
(Note: main effect is Mn)
tHonq et al. (2015)
5 administrative regions
South Korea
NR
Cross-sectional
n: 1001
General population of
children
Blood
Venous blood; GFAAS
Age at Measurement:
8-11 yr old
Median: 1.81 pg/dL
75th: 2.25 pg/dL,
95th: 3.01 pg/dL
Max: 6.16 pg/dL.
KEDI-WISC
Age at outcome:
8-11 yr old
Age, sex, residential
region, paternal education
level, and yearly income
Iog10-transformed blood
Hg, Mn, urine
concentrations of cotinine,
phthalate metabolites
B=-1.948 (-3.608,
-0.288), adjusted for
other metals
B=-2.113 (-3.73,
-0.496), adjusted for
ADHD and CPT
B=-2.118 (-3.792,
-0.445), adjusted for
socio-demographic
factors
tMenezes-Filho et al.
(2018)
Salvador, Bahia
Brazil
Cross-sectional
Study years: NR
School-based cohort
n: 225
Children from 4
elementary schools in
industrial town.
Blood
Child venous blood Pb,
hair and toenails tested
for Mn; GFAAS
Age at Measurement:
7-12 yr
Mean: 1.64 pg/dL,
Median: 1.15 pg/dL, only
about 2% of children
above the Centers for
Disease Control and
Prevention ref value of 5
pg/dL
75th: 2.1 pg/dL
Max: 15.6 pg/dL.
IQ estimated using
vocabulary and matrix
reasoning (WASI).
Age at outcome:
7-12 yr
Age, Maternal IQ
Effect modification by Mn
assessed. Both Pb and
Mn were log-transformed
to include in the model.
Interaction between Pb
and Mn assessed.
Put results for Model B
which is adjusted for Mn,
age and maternal IQ
3-274
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tLucchini et al. (2012)
Junior high school-age
Blood
IQ tested using WISC-III
Sex, age at testing,
B=-2.248 (-4.111,
children from 20 local
(verbal IQ and
parental education, SES,
-0.385)
Valcamonica and Garda
public schools
Children's venous blood;
performance IQ
family size, parity order,
Lake areas in Province of
n: 299
GFAAS
assessed).
BMI
Brescia
Italy
Age at Measurement:
Age at outcome:
Cross-sectional
11-14 yr
11-14 yr
1.71 pg/dL, Median: 1.50
75th: 2.10 pg/dL
Max: 10.2 pg/dL.
AAS = atomic absorption spectrometry; ALSPAC = Avon Longitudinal Study of Parents and Children; avg = average; BLL = blood lead level; BNT = Boston Naming Test; CARES =
Communities Actively Researching Exposure Study; CCAAPS = Cincinnati Childhood Allergy and Air Pollution Study; Cd = cadmium; CI = confidence interval; CKD = chronic kidney
disease; CKiD = Chronic Kidney Disease in Children Study; ETS = environmental tobacco smoke; FSIQ = full-scale intelligence quotient; GFAAS = graphite furnace atomic absorption
spectrometry; GM = geometric mean; Hg = mercury; HOME = Health Outcomes and Measures of the Environment; ICP-MS = inductively coupled plasma mass spectrometry; IQ =
intelligence quotient; KEDI = Korean Educational Development Institute; MIREC = Maternal-Infant Research on Environmental Chemical; Mn = manganese; mo = month(s); NA = not
available; NR = not reported; Pb = lead; PC = primary caregiver; PIQ = performance IQ; Q = quartile; SD = standard deviation; SES = socioeconomic status; VIF = variance inflation
factor; VIQ = verbal IQ; WASI = Wechsler Abbreviated Scale of Intelligence; WISC = Weschler Intelligence Scale for Children; wk = week(s); WPPSI = Wechsler Preschool and
Primary Scale of Intelligence; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResults are not standardized (e.g., BLL distribution data needed to calculate the standardized estimate was not reported or categorical data was analyzed).
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-275
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Table 3-3E Epidemiologic studies of Pb exposure and infant development
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Bellinger et al. (1987)
Boston, MA
U.S.
Apr. 1979 - Apr. 1981
(enrollment)
Followed through 2 yr
Cohort
Birth cohort, n = 182
infants
Recruitment from births
at Brigham and Women's
Hospital
Blood
Cord blood; anodic stripping
voltammetry (ASV)
Age at measurement:
Delivery
Mean (SD): 6.6 (3.2) pg/dL
Low: <3 pg/dL
Medium: 6-7 pg/dL
High: >10 pg/dL
MDI assessed using
BSID-II
Age-standardized
scores (mean: 100,
SD: 16)
Age at outcome: 2 yr
Maternal age, race, IQ,
education, years of
smoking, and alcohol
drinks/wk in 3rd
trimester, SES, HOME
score, child sex, birth
weight, gestational age,
birth order.
Beta:
Cord blood:
Low vs. high: -3.8 (-6.3,
-1.3)
Medium vs. high: -4.8
(-7.3, -2.3)
Concurrent blood
reported not to be
associated with MDI,
quantitative data not
reported.
Jedrvchowski et al.
(2009b)
Krakow, Poland
2001-2004 (enrollment)
Followed through 3 yr
Cohort
Birth cohort, n = 381-415 Blood
children
Recruited pregnant
mothers from prenatal
clinics in Krakow inner
city in 1st and 2nd
trimesters.
Cord blood; ICP-MS
Age at measurement:
Delivery
GM (95% CI): 1.29 (1.24,
1.34) pg/dL
Median: 1.23 pg/dL
MDI assessed using
BSID-II (Polish
version)
Standardized scores
Age at outcome:
12, 24, 36 mo
Maternal education and
prenatal smoking, child
sex and birth order
Beta
Age 2 yr: -1.8 (-3.4,
-0.14)
Age 3 yr:
-0.21)
-1.6 (-2.9,
Henn et al. (2012)
Mexico City
Mexico
1997-2000 (enrollment)
Followed through 24 mo
N: 455
Blood
MDI assessed using Sex, gestational age,
Women recruited during Child venous blood; ICP-MS
pregnancy or at delivery
Age at measurement:
12, 24 mo
Mean (SD):
BSID-II (Spanish
version)
Age at Outcome:
12, 18, 24, 30, 36
mo
hemoglobin, maternal IQ,
maternal education, and
visit
Beta
12-months: -0.07 (-0.39,
0.25)
24-months: -0.08 (-0.46,
0.30)
12 mo Pb * Mn <2 pg/dL:
-0.31 (-1.25, 0.62)
3-276
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Cohort
12 mo: 5.1 (2.6) pg/dL
24 mo: 5.0 (2.9) pg/dL
24 mo Pb * Mn <2 pg/dL:
-1.27 (-2.18, -0.37)
Hu et al. (2006)
N: 83 (cord) - 146 (24-
Blood
MDI assessed using
Maternal age and IQ,
Maternal T1: -0.76
months child blood)
BSID-II (Spanish
child sex, current weight,
(-1.50, -0.03)
Mexico City
Mexico
Women recruited during
Maternal blood, cord blood,
and child venous blood; ICP-
version)
height-for-age Z score,
and concurrent blood Pb
(in models examining
prenatal blood Pb)
Maternal T3: -0.43
(-1.10, 0.27)
pregnancy or at delivery
MS
Age at Outcome:
Cord: -0.06 (-0.87, 0.74)
1997-2000 (enrollment)
Followed through 24 mo
Age at measurement: T1, T2,
T3 (maternal), delivery (cord),
12, 24 mo (child)
24 mo
Child 24 mo: -0.23
(-0.92, 0.45)
Cohort
Mean (SD):
Maternal T1: 7.07 (5.10)
|jg/dL; T3: 6.86 (4.23) pg/dL
Cord: 6.20 (3.88) pg/dL
Child 24 mo: 4.79 (3.71)
pg/dL
tY Ortiz etal. (2017)
Mexico City
Mexico
Jul 2007-Feb 2011
Followed through 24 mo
Cohort
PROGRESS birth cohort
n: 536
Women <20 wk of
gestation and planning to
reside in Mexico City for
the next 3 yr.
Blood
Maternal blood; ICP-MS.
Age at measurement:
T2, T3
Mean:
T2: 3.7 pg/dL
T3: 3.9 pg/dL.
Cognitive and
language
development
assessed using
BSID-III.
Standardized scores
(mean: 100, SD:
15). Cognitive,
language and motor
scores were jointly
considered for
standardizing.
Age at outcome:
24 mo
Infant sex, birth weight,
gestational age, maternal
age, maternal IQ (WAIS
Spanish version), HOME
score.
Beta
Cognitive
Development:
T2: 0.76 (-3.35, 4.87)bc
T3: -6.60 (-13.49,
0.29)bc
Stress2: -0.23 (-0.45,
-0.01 )bc
T3*Stress: 1.02 (-0.78,
2.82)bc
Language
Development:
T2: 0.97 (-3.18, 5.12)bc
T3: -6.00 (-12.94,
0.94)bc
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Outcome
Confounders
Effect Estimates and
95% Clsa
tKimetal. (2013c) and MOCEH study
Kimet al. (2013b)
Seoul,
Ulsan
Korea
Cheonan and
2006-2010
Followed through 6 mo
Cohort
n: 884
Mothers recruited before
20th wk of pregnancy
between and were in
locations (Seoul,
Cheonan and Ulsan).
Blood
Maternal venous blood;
GFAAS with Zeeman
background correction,
measured for Pb and Cd
Age at measurement:
Early (<20 wk) and late
pregnancy (med = 39 wk)
Early pregnancy: 1.4 (GM),
2.1 (90th), 9.8 (max) pg/dL
Late pregnancy: 1.3 (GM); 2.1
(90th), 4.3 (max) pg/dL
MDI assessed using
BSID-II (Korean
version)
Age-standardized
scores (mean: 100,
SD: 15).
Age at outcome:
6 mo
Birth weight, infant sex,
maternal age and
education, family income,
breastfeeding status,
residential area.
Beta
Early:
Overall: 0.02 (-1.20,
1.24)
Cd <1.47 pg/L: 2.44
(0.04, 4.83)
Cd >1.47 pg/L: -0.87
(-2.52, 0.78)
Late:
Overall: -1.74 (-3.37,
-0.12)
Cd <1.51 pg/L: -0.29
(-2.88, 2.30)
Cd >1.51 pg/L: -3.20
(-5.35, -1.06)
tKimetal. (2018b)
4 cities: Seoul, Anyang,
Ansan and Jeju
Korea
2011-2012 (enrollment)
Followed through 24 mo
Cohort
CHECK cohort
n: 140
birth cohort- pregnant
women recruited from 4
cities in Korea before
delivery.
Blood
Maternal and cord blood;
method NR
Age at measurement:
Delivery
Median (IQR):
Maternal: 2.7 (3.5, 5.7) pg/dL
Cord: 1.2 (0.8, 1.7) pg/dL
MDI assessed using
BSID-II (Korean
version)
Age at outcome:
13-24 mo
BPA, and phthalates,
maternal age
(continuous), birth
delivery mode
(categorical), monthly
household income
(categorical), child's sex,
and BDI (continuous) of
the mother, gestational
age (continuous),
primiparous (categorical),
and
pre-pregnancy BMI
(categorical).
Associations of blood Pb
concentrations and MDI
were assessed but not
reported because they
lacked statistical
significance.
tValeri et al. (2017)
Birth cohort
Blood
Cognitive and
Child sex, age attesting,
Beta
n: 825 (Pabna: 409,
language
maternal age and
Pabna
Pabna and Sirajdikhan
Sirajdikhan: 416)
Cord blood; ICP-MS,
development using
education, maternal IQ,
Cognitive: 0.012 (-0.05,
0.074)
districts
measured for Pb, As, and Mn
BSID-III (Bengali
HOME score, ETS,
Bangladesh
Mother-infant pairs,
version, adapted for
protein intake.
2010-2013 (enrollment)
Age at measurement:
rural Bangaldesh)
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Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Followed through 20-40
mo
Cohort
pregnant women enrolled Delivery
in the first trimester.
GM: Pabna: 1.8 |jg/dL,
Sirajdikhan: 6.0 |jg/dL.
75th: Pabna: 2.4, Sirajdikhan:
9.7
Max: Pabna: 79.1 |jg/dL,
Sirajdikhan: 36.0 |jg/dL.
Two primary
outcomes derived by
summing across raw
scores of cognitive
and language
development. Z-
scores were
calculated.
Age at outcome:
20-40 mo
Language: -0.014
(-0.076, 0.048)
Sirajdikhan
Cognitive: -0.011
(-0.024, 0.001)
Language: -0.004
(-0.026, 0.017)
tKoshv et al. (2020)
Old Town, Salavanpet
and neighboring areas in
Vellore, South India
Mar 2010-Feb 2012
(enrollment)
Followed through 5 yr
Cohort
Etiology, Risk Factors
and Interactions of
Enteric Infections and
Malnutrition and the
Consequences for Child
Health and Development
(MAL-ED) Network
n: 228 (followed 2 yr) and
212 children (followed 5
yr)
Birth cohort of mother-
infant pairs in eight
adjacent urban slum
dwelling areas
Blood
Child venous blood; GFAAS.
Mean BLL derived by
averaging BLLs at 15 and 24
mo for analysis at 2 yr, and
15, 24 and 36 mo for analysis
at 5 yr.
Age at measurement:
7, 15, 24 and 36 mo
Mean: 15 mo: 0.5 pmol/L, 24
mo: 0.6 pmol/L, 36 mo: 0.6
pmol/L
Cognitive and
language
development
assessed using
BSID-III (culturally
adapted and
translated)
Raw scores of
cognition and
expressive and
receptive language
domains.
Age at outcome:
24 mo
Child sex, maternal
intelligence raw scores,
SES, mean body Fe
levels.
Beta
0.2,
Cognitive: -0.2
-0.03)
Expressive language:
-0.2 (-0.3, -0.1)
Receptive language:
-0.04 (-0.1, 0.02)
tShekhawat et al. (2021) n:117
Western Rajasthan
India
2018-2019 (enrollment)
Followed through 6.5 mo
(average)
Mother-child pairs in third
trimester or at delivery
Blood
Cord blood; ICP-OES
Age at measurement:
Delivery
GM =4.14 |jg/dL; mean =
4.77 ± 3.3 |jg/dL; median =
Cognitive and
language
development
assessed using
BSID-III
Age at outcome: 6.5
mo (average)
Maternal age, gravida,
gestational age, maternal
education, child sex and
weight, preterm birth,
maternal food intake
during pregnancy,
smoking, alcohol
consumption, maternal
residential and
13 (95 % CI)
(A) Umbilical cord Pb
level <5 |jg/dL (n = 70)
Composite cognitive:
0.19 (-0.03, 0.34)
Composite language:
0.21 (-0.23, 0.42)
Subscale receptive
language: 0.11 (-0.6,
1.66)
Subscale expressive
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Outcome
Confounders
Effect Estimates and
95% Clsa
Cohort
4.23 |jg/dL
75th: 5.1 |jg/dL.
occupational history,
delivery type.
language: 0.22 (-0.03,
1.87)
(B) Umbilical cord Pb
level = 5.0-10.5 |jg/dL (n
= 47)
Composite cognitive:
-0.13 (-0.77, 0.28)
Composite language:
-0.05 (-0.7, 0.47)
Subscale receptive
language: -0.04 (-3.5,
2.5)
Subscale expressive
language: 0.04 (-3.8,
2.9)
tParaiuli et al. (2015a) Birth cohort from
Bharatpur General
Hospital
n: 100
Chitwan, Bharatpur
District
Nepal
Sep-Oct 2008
(enrollment)
Followed through 24 mo
Resided in area for at
least 2 yr delivered at
term (i.e., >37 wk).
Blood
Cord blood; ICP-MS,
measured for Pb, As and Zn
Age at measurement:
Delivery
Median: 2.06 |jg/dL
Max: 22.08 pg/dL.
MDI assessed using
BSID-II
Age at outcome:
24 mo
Maternal age and
education, BMI,
gestational age, family
income, parity, birth
weight, weight at 24 mo,
child age assessment,
As, Zn, HOME score
(smoking and alcohol
consumption not included
given low prevalence).
Beta
-4.21
-13.62, 5.20)c
Cohort
tParaiuli et al. (2015b)
Chitwan, Bharatpur
district
Nepal
Sep-Oct 2008
(enrollment)
Followed through 36 mo
Cohort
Birth cohort from
Bharatpur General
Hospital
n: 100
Resided in area for at
least 2 yr delivered at
term (i.e., >37 wk).
Blood
Cord blood; ICP-MS,
measured for Pb, As and Zn
Age at measurement:
Delivery
Median: 2.06 pg/dL
Max: 22.08 pg/dL.
MDI assessed using Maternal age and
BSID-II
Age at outcome:
36 mo
education, BMI,
gestational age, family
income, parity, birth
weight, weight at 24 mo,
child age at assessment,
As, Zn, HOME score
(smoking and alcohol
consumption not included
given low given low
prevalence).
Beta
4.05 (-3.21, 11.311
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Design
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Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
t Zhou et al. (2017)
Shanghai
China
2010-2012 (enrollment)
Followed through 24-36
mo
Cohort
Shanghai Stress Birth
Cohort Study
n: 139
Women enrolled in
prenatal clinics of
maternity hospitals
during mid-to-late
pregnancy.
Blood
Maternal blood; AAS
Age at measurement:
28-36 wk of gestation
GM (95% CI): 3.30 (3.05,
3.57) |jg/dL.
Language
development
assessed using
GDS (Chinese
version)
Age at outcome:
24-36 mo
Maternal age at
enrollment, economic
status, maternal
education, gestational
week, child sex, birth
weight and age.
Beta per log—10
transformed BLL
Language development
Overall: -6.76 (-17.29,
3.77)d
Low stress: -1.76
(-13.03, 9.51 )d
High stress: -33.82
(-60.04, —7.59)d
tViqeh et al. (2014)
Tehran
Iran
October 2006 - March
2011
Followed through 36 mo
Cohort
Birth cohort
n: 174
Mother-infant pairs
recruited in first trimester
(8-12 wk).
Blood
Maternal blood,
ICP-MS
cord blood;
Age at measurement:
3 trimesters during pregnancy
and delivery
Mean:
Maternal T1: 4.15 pg/dL, T2:
3.44 pg/dL, T3: 3.78 pg/dL
Cord: 2.86 pg/dL
Max:
Maternal T1: 20.5 pg/dL, T2:
7.5 pg/dL, T3: 8.0 pg/dL
Cord: 6.9 pg/dL
Mental development
composite assessed
using the ECDI by
Harold Ireton
(language
comprehension,
expressive
language, gross
motor, self-help,
social interaction).
Cutoff point scores
for development
delay were score
<20% of that
expected for
children's age.
Age at outcome:
36 mo
Maternal educational,
BMI, family income,
gestational age, birth
weight, birth order (first
born).
OR
Total ECDI: 1.74 (1.1?
2.5).
tLin et al. (2013)
Taipei, Taiwan
April 2004-Jan 2005
(enrollment)
Birth cohort
n: 230
Mother-infant pairs from
medical center, local
hospital, and obstetric
clinics.
Blood
Maternal blood, cord blood;
ICP-MS, measured for Pb,
Mn, As, and Hg.
Cognitive and
language
development
assessed using
CDIIT
Maternal age, education,
infant gender, ETS
during pregnancy and
after delivery, fish intake,
and HOME score.
Beta
Cognitive
High vs. Low Pb: -5.35
(-9.642, —1.058)b
High Mn*low Pb: -4.15
(-9.618, 1.318)b
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Design
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Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Followed through 2 yr
Panel Study
Pb categories:
Low: <16.45 |jg/L
High: >16.45 pg/L
Mn categories:
Low: <59.59 |jg/L
High: >59.59 pg/L
Age at measurement:
delivery
Mean: 13 pg/L, GM: 10.61
pg/L
75th: 16.45 pg/L
Max: 43.22 pg/L
Age at outcome:
2 yr
Low Mn*high Pb: -4.79
(-10.298, 0.718)b
High Mn*high Pb: -8.19
(-14.403, —1.977)b
Language
High vs. Low Pb: -2.53
(-6.234, 1.174)b
High Mn*low Pb: -1.56
(-6.264, 3.144)b
Low Mn*high Pb: 0.22
(-4.523, 4.963)b
High Mn*high Pb: -6.81
(-12.161, -1.459)b
tNozadi etal. (2021)
Navajo Nation
United States
February 2013-June
2018 (enrollment)
Followed through 10-13
mo
Cohort
Navajo Birth Cohort
Study
n: 327
Blood
Maternal blood, child blood;
ICP-DRC-MS.
Age at measurement:
Delivery or 36-wk visit
(maternal); 10, 13 mo (child)
GM = 0.410 pg/dL; median =
0.37 pg/dL
75th: 0.51 pg/dL
95th: 1.20 pg/dL
Problem-solving
scores assessed
using ASQ:I.
Age-adjusted
scores.
Age at outcome: 10-
13 mo
Urine strontium and
arsenic.
Beta
Problem-Solving: -0.67
(-1.54, 0.20)
tNvanza et al. (2021)
Northern Tanzania
Tanzania
2015-2017 (enrollment)
Followed through 6-12
mo
Mining and Health
Prospective Longitudinal
Study in Northern
Tanzania
n: 439
Birth cohort of mother-
child pairs recruited in
2nd trimester
Maternal dried blood spots;
ICP-MS, measured for Pb,
Hg, and Cd
Age at measurement:
T2
Median: 2.72 pg/dL
Language and
global
neurodevelopment
assessed using
MDAT. Scores in
each domain
classified as normal
(>90th percentile on
all items in that
Maternal age and
education, maternal and
paternal occupation,
number siblings under 5
yr at home, and family
SES, infant sex, age,
birth weight, height, and
weight as a proxy for
nutritional status.
Prevalence ratio
Language Development:
1.0 (1.0, 1.0)
Global
neurodevelopmental
status: 1.0 (0.9, 1.0)
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Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Cohort
75th: 4.25 pg/dL
Max: 14.5 pg/dL
domain or <90th
percentile on one or
two items in the
domain) or impaired
(<90th percentile on
more than two items
in a domain).
Age at outcome:
6-12 mo
(Covariates with p < 0.20
retained in the final
models.)
Hg >0.08 pg/dL * Pb> 3.5
pg/dL: 1.4 (0.9, 2.1)
AAS = atomic absorption spectrometry; As = arsenic; ASQ:i = Ages and Stages Questionnaires: Inventory; BDI = Beck Depression Inventory; BLL = blood lead level; BMI = body
mass index; BPA = bisphenol A; BSID = Bayley Scales of Infant and Toddler Development; CARES = Communities Actively Researching Exposure Study; CCAAPS = Cincinnati
Childhood Allergy and Air Pollution Study; Cd = cadmium; CDIIT = Comprehensive Developmental Inventory for Infants and Toddlers; CHECK = Children's Health and Environmental
Chemicals in Korea; CI = confidence interval; CKD = chronic kidney disease; CKiD = Chronic Kidney Disease in Children Study; ECDI = Early Child Development Inventory;
ELEMENT = Early Life Exposure in Mexico to Environmental Toxicants; ETS = environmental tobacco smoke; FSIQ = full-scale IQ; GDS = Gesell Developmental Schedules; GFAAS
= graphite furnace atomic absorption spectrometry; GM = geometric mean; Hg = mercury; HOME = Health Outcomes and Measures of the Environment; ICP-DRC-MS = dynamic
reaction cell for inductively coupled plasma mass spectrometry; ICP-MS = inductively coupled plasma mass spectrometry; ICP-OES = inductively coupled plasma optical emission
spectrometry; IQ = intelligence quotient; MDAT = Malawi Development Assessment Tool; MDI = Mental Developmental Index; Mn = manganese; mo = month(s); MOCEH = Mothers'
and Children's Environmental Health; NBAS = Neonatal Behavioral Assessment Scale; NR = not reported; OR = odds ratio; Pb = lead; PDI = Psychomotor Developmental Index;
PROGRESS = Programming Research in Obesity, Growth, Environment and Social Stressors; SD = standard deviation; SES = socioeconomic status; T1 = first trimester of
pregnancy; T2 = second trimester of pregnancy; T3 = third trimester of pregnancy; WISC = Weschler Intelligence Scale for Children; wk = week(s); yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
The CI was calculated from a p-value and the true CI may be wider or narrower than calculated.
°Results are unstandardized because the log base used for exposure transformation was unspecified in the study.
dResults are unstandardized because the Pb level distribution data was not available.
tStudies published since the 2013 Integrated Science Assessment for Lead.
Table 3-4E Epidemiologic studies of Pb exposure and performance on neuropsychological tests of cognitive
function, i.e., learning, memory, and executive function
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Lanphear et al. (2000)
United States
U.S. NHANES Blood
n = 4,853 children ages
6-16 yr (born 1972-1988) Concurrent
GM (SD): 1.9 (7.0)
Digit span
WISC-R
Age at outcome: 6-16 yr
Child sex, race/ethnicity, -0.05 (-0.09, -0.01)
poverty index ratio,
reference adult education,
serum ferritin and cotinine
levels. Did not consider
potential confounding by
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Design
Study Population Exposure Assessment
Outcome
Effect Estimates and
Confounders 950/o C|sa
1988-1994
Cross-sectional
Large U.S. representative 63.5% <2.5
study of multiple risk Detection limit = 0.5
factors and outcomes , x , , , „
Interval analyzed: 1-5
Linear regression
parental cognitive function
or caregiving quality.
Kriea et al. (2010)
United States
1991-1994
Cross-sectional
U.S. NHANES
n = 773 children ages 12-
16 yr (born 1972-1982)
Concurrent
GM (SD): 1.9 (7.0)
63.5% <2.5
Detection limit = 0.5
Large U.S. representative , x , , , „
study of multiple risk lnterval analyzed: 1-5
factors and outcomes
Digit span
WISC-R
Age at outcome: 6-16 yr
Log-linear regression
Child sex, caregiver
education, family income,
race/ethnicity, test
language. Did not
consider potential
confounding by parental
cognitive function or
caregiving quality.
-0.34 (-0.59, -0.08)
Surkan et al. (2007)
Boston, Massachusetts
and Farmington, Maine
United States
Cross-sectional
n = 389 children
6-10 yr
Recruitment from trial of
amalgam fillings
Blood
Concurrent
Group 1: 1-2
Group 2: 3-4
Group 3: 5-10
Mean (SD): 2.2 (1.6)
General memory index,
WRAML
Age at outcome: 6-10 yr
Caregiver IQ, child age,
SES, race, birth weight.
Also considered potential
confounding by site, sex,
birth order, caregiver
education and marital
status, parenting stress,
and maternal utilization of
prenatal and annual
health care but not
parental caregiving
quality.
-0.69 (-4.4, 3.0)
-6.7 (-12, -1.2)
tYorifuii et al. (2011)
Faroese island
Denmark
1986-1987 (enrollment)
Followed through 7-14 yr
Cohort
Birth cohort
n: At age 7: 896, At age
14: 808
Birth cohort of mother-
infant pairs
Blood, hair
Cord blood;
electrothermal AAS.
Age at measurement:
At birth
GM of cord blood Pb:
1.57 |jg/dL
75th: 2.2 pg/dL
Verbal and visuospatial
reasoning, language,
learning, and memory
assessed using WISC-R
similarities, WISC-R block
designs, BNT, and CVLT-
C.
Age at outcome:
7, 14 yr
Age, sex, maternal
Raven's score, paternal
employment and
education, maternal
education, daycare at age
7, medical risk, and
maternal alcohol use and
smoking during
pregnancy
Beta
WISC-R at 7 yr old with
cord mercury
Block Design: -0.011
(-0.083, 0.062)
Similarities: -0.122
(-0.38, 0.135)
Digit Span Forward: -0.1
(-0.183, -0.016)
WISC-R at 7 yr old
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Outcome
Confounders
Effect Estimates and
95% Clsa
Block Design: -0.004
(-0.013, 0.006)
Similarities: 0.019
(-0.051, 0.088)
Digit Span Forward:
-0.028 (-0.049, -0.006)
CVLT-C at 14 yr old plus
interaction with cord
mercury
Recognition: -0.053
(-0.136, 0.031)
Long-term Recall: -0.1
(-0.256, 0.057)
Short-term Recall: -0.009
(-0.178, 0.16)
Learning: -0.438 (-0.965,
0.089)
CVLT-C at 14 yr old
Recognition: -0.001
(-0.022, 0.021)
Long-term Recall: 0.041
(0, 0.082)
Short-term Recall: 0.013
(-0.031, 0.058)
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Reference^and Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Learning: 0.037 (-0.103,
0.177)
Boston Naming Test at 14
yr old plus interaction with
cord mercury
With Cues: 0.002 (-0.337,
0.342)
No Cues: -0.095 (-0.473,
0.283)
Boston Naming Test at 14
yrold
With Cues: 0.033 (-0.056,
0.122)
No Cues: 0.003 (-0.096,
0.102)
WISC-R at 14 yr old plus
interaction with cord
mercury
Block Design: 0.241
(-0.575, 1.057)
Similarities: -0.147
(-0.383, 0.089)
Digit Span Backward:
-0.16 (-0.253, -0.067)
Digit Span Forward:
-0.107 (-0.199, -0.016)
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Reference^and Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Digit Span: -0.267
(-0.423, -0.111)
WISC-R at 14 yr old
Block Design: -0.023
(-0.238, 0.191)
Similarities: -0.007
(-0.069, 0.055)
Digit Span Backward:
-0.035 (-0.06, -0.009)
Digit Span Forward:
-0.024 (-0.048, 0)
Digit Span: -0.059 (-0.1,
-0.017)
CVLT-C at 7 yr old plus
interaction with cord
mercury
Recognition: -0.094
(-0.2, 0.011)
Long-term Recall: 0.037
(-0.141, 0.215)
Short-term Recall: -0.068
(-0.22, 0.084)
Learning: -0.501 (-0.981,
-0.02)
CVLT-C at 7 yr old
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Reference^and Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Recognition: -0.003
(-0.032, 0.026)
Long-term Recall: 0.033
(-0.014, 0.081)
Short-term Recall: 0.043
(0.003, 0.084)
Learning: 0.073 (-0.057,
0.202)
Boston Naming Test at 7
yr old plus interaction with
cord mercury
With Cues: -0.138
(-0.469, 0.193)
No Cues: -0.046 (-0.376,
0.284)
Boston Naming Test at 7
yrold
With Cues: 0.042 (-0.045,
0.129)
No Cues: 0.039 (-0.049,
0.127)
tTatsuta et al. (2014)
Sendai, Tohoku region
Japan
Study years NR
Followed through 42 mo
TSCD birth cohort
n: 387
Mother-infant pairs urban
areas of the Tohoku
district
Blood
Cord blood; ICP-MS.
Age at measurement:
Delivery
Intelligence and
achievement (K-ABC)
Age at outcome:
42 mo
Child sex, birth order,
alcohol and smoking
habits, duration of
breastfeeding, annual
family income at 42 mo,
and maternal IQ (Raven
SPM)
Beta
Adjusted Model
Mental Processing Score:
-3.319 (-12.41, 5.774)
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Outcome
Confounders
Effect Estimates and
95% Clsa
Cohort
Median: 1.0 pg/dL
Max: 1.8 pg/dL
Sequential Processing
Score: -2.136 (-12.80,
8.531)
tOppenheimer et al.
(2022)
New Bedford Harbor,
Massachusetts
United States
1993-1998 (enrollment)
Followed through 2008-
2014
Cohort
New Bedford Cohort
n: 373
Blood
Cord blood; isotope
dilution ICP-MS
Age at measurement:
Delivery
Mean (SD): 1.4 (0.9)
pg/dL
Max: 9.4 |jg/dL
Cognitive Effects
Cognitive effects were
assessed using four
subtests of the Delis-
Kaplan Executive
Function System. These
included Trail Making:
Number-Letter Switching
condition, Verbal Fluency:
Category Switching
condition, Design
Fluency: Filled Dots and
Empty Dots Switching
condition, and Color-Word
Interference:
Inhibition/Switching
condition.
Multiple linear regression
models adjusted for child
race, sex, age at exam,
year of birth, HOME
score, maternal marital
status at child's birth,
maternal IQ, maternal
seafood consumption
during pregnancy,
maternal smoking during
pregnancy, maternal and
paternal education, and
annual household income
at child's birth, and study
examiner.
Beta
WRAML
Verbal Working Memory:
0.12 (-0.20, 0.45)
Symbolic Working
Memory: 0.09 (-0.246,
0.42)
Working Memory Index
Differences: 0.59 (-0.97,
2.15)
Age at Outcome:
14-18 yr
Choet al. (2010)
Seoul (metropolitan),
Seongnam (suburban),
Ulsan and Incheon
(industrial), and
Yeoncheon (rural)
South Korea
2009
Cross-sectional
n = 639 children (8-11 yr) Blood
Color-Word score
School-based recruitment Child blood; GFAAS with SCWT
Zeeman background
correction
Age at measurement:
8-11 yr
Mean (SD): 1.9 (0.67)
pg/dL
10th—90th: 1.2-2.8 pg/dL
Age at outcome:
8-11 yr
Age, sex, paternal
education, maternal IQ,
child IQ, birth weight,
urinary cotinine,
residential area. Did not
consider potential
confounding by parental
caregiving quality.
Beta
0 (-0.02, 0.02)
3-289
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Reference^and Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tFruh etal. (2019)
Eastern Massachusetts
U.S.
1999-2002 (enrollment)
Followed through 7 yr
Cohort
Project Viva
n: 1006
Birth cohort of mother-
child pairs
Blood
Maternal venous
erythrocyte blood
specimens; ICP-MS
Age at measurement:
2nd to 3rd trimester of
pregnancy (median: 27.9
wk)
Med (IQR): 1.1 (0.06)
pg/dL
Executive Function (see
also Section 3.5.1)
Parent teacher ratings on
BRIEF
Age at outcome:
7 yr
Scores standardized for
child age and sex;
Additional adjustment for
maternal 2nd trimester Hg
and Mn levels, nulliparity,
smoking during
pregnancy, IQ, and
education; Paternal
education; HOME
composite score and
household income; and
child race/ethnicity.
Beta
BRIEF Parent-Reported
Behavioral Regulation
Index
All: 1.15 (-0.217, 2.517)
Girls: 1.717 (0.025, 3.408)
Boys: 0.85 (-1.058,
2.758)
Metacognition Index
All: 0.95 (-0.25, 2.15)
Girls: 1.483 (-0.108,
3.075)
Boys: 0.6 (-1.05, 2.25)
General Executive
Composite
All: 1.217 (-0.1, 2.533)
Girls: 1.95 (0.1, 3.8)
Boys: 0.783 (-0.958,
2.525)
BRIEF Teacher-
Reported
Behavioral Regulation
Index
All: 0.767 (-0.567, 2.1)
3-290
-------
Reference^and Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Girls: 0.933 (-1.142,
3.008)
Boys: 0.75 (-0.9, 2.4)
Metacognition Index
All: 0.683 (-0.717, 2.083)
Girls: 0.9 (-1.308, 3.108)
Boys: 0.683 (-1.142,
2.508)
General Executive
Composite
All: 0.7 (-0.65, 2.05)
Girls: 0.883 (-1.258,
3.025)
Boys: 0.683 (-1.008,
2.375)
tFruh etal. (2021)
Eastern Massachusetts,
U.S.
1999-2002 (enrollment),
Followed through 6-11 yr
Cohort
Project Viva
n: 2128
Birth cohort of mother-
child pairs
Blood
Maternal venous
erythrocyte blood
specimens; ICP-MS
Age at measurement:
2nd to 3rd trimester of
pregnancy (median: 27.9
wk)
Global Executive
Composite (GEC);
Difficulties score
BRIEF; SDQ
Age at Outcome:
6-11 yr
Scores on BRIEF
Total standardized forage, sex
and model adjusted for
child race/ethnicity,
maternal parity, maternal
smoking status, maternal
IQ, maternal education,
paternal; education,
hemoglobin, HOME
score, household income,
and fish consumption.
SDQ model adjusted for
age, sex, and above
covariates. Pb, Mn, Se
Beta
SDQ Total Difficulties:
0.617 (-0.058, 1.292)
BRIEF GEC:
1.11 (-0.12, 2.34)
3-291
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Med (IQR): 1.1 (0.06)
pg/dL
and MeHg included
together in the models.
tRuebner et al. (2019)
46 centers
U.S.
Study Years: NR
Followed through: 1-16 yr
Cohort
CKiD Cohort study
n: 412
Children with mild to
moderate CKD
Blood
Child venous blood; ICP-
MS. The BLL
measurement closest to
the time of neurocognitive
testing was used for
analysis (concurrent).
Age at measurement:
NR; 2, 4, or 6 yr after
study entry
Median: 1.2 |jg/dL
75th: 1.8 |jg/dL
Max: 5.1 |jg/dL
Executive function (see
also Section 3.5.1 [FSIQ],
Section 3.5.2 [attention
and hyperactivity])
Age-specific assessments
administered at visit 3, 5,
7, or 9. Last available
results used (mean time
between BLL and
outcome assessment =
2.3 yr). Delis-Kaplan
Executive Function
System Tower Subset (>6
yr), BRIEF-P (2-5 yr),
BRIEF (6-18), BRIEF-A
(>18 yr)
Age at outcome:
1 to >18 yr
Age, sex, race, poverty,
maternal education.
Adjusted BRIEF results
were not reported
because they were not
statistically significant.
tMerced-Nieves et al.
(2022)
Mexico City
Mexico
2007-2011 (enrollment)
Followed through 6-7 yr
Cohort
PROGRESS Cohort
n = 549
Birth cohort
Blood
Maternal, cord, and child
blood; Agilent 8800 ICP
Triple Quad
Age at measurement:
Maternal: T2, T3, delivery
Cord: delivery
Postnatal: 4-6 yr
Mean (SD)
Maternal T2: 2.7 (2.7)
pg/dL
Various measures from
Condition Position
Responding (CPR),
Temporal Response
Differentiation (TRD),
Delayed Matching-to-
Sample (DMTS), and
Incremental Repeated
Acquisition (IRA) from the
OTB
Age at Outcome:
6-7 yr
Child's age at testing, Betas for BLL at T3
maternal education (high school), and SES
Modification by sex
examined.
Observing response
latency:
0.001s (-0.08, 0.08s)
TRD
Average latency:
0.14s (-0.001, 0.29s)
DMTS
Average observing
response latency:
0.08s (-0.04, 0.20s)
3-292
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Maternal T3: 3.9 (2.8)
pg/dL
Maternal at delivery: 4.3
(3.2) pg/dL
Cord: 3.4 (2.6) pg/dL
Child: 2.4 (2.6) pg/dL
IRA
Effective response rate:
-0.01s (-0.03, -0.002s)
AAS = atomic absorption spectrometry; BLL = blood lead level; BNT = Boston Naming Test; BRIEF = Behavior Rating Inventory of Executive Functions; CANTAB = Cambridge
Neuropsychological Test Automated Battery; CI = confidence interval; CKD = chronic kidney disease; CKiD = Chronic Kidney Disease in Children Study; CVLT-C = California Verbal
Learning Test-Children's version; ETS = environmental tobacco smoke; FSIQ = full-scale IQ; GFAAS = graphite furnace atomic absorption spectrometry; GM = geometric mean;
HOME = Health Outcomes and Measures of the Environment; ICP-MS = inductively coupled plasma mass spectrometry; K-ABC = Kaufman Assessment Battery for Children; MANAs
= Metals, Arsenic and Nutrition in Adolescents Study; MeHg = methyl mercury; NHANES = National Health and Nutrition Examination Survey; NR = not reported; OTB = Operant Test
Battery; Pb = lead; SCWT = Stroop Color-Word test; SD = standard deviation; SDQ = Strengths and Difficulties Questionnaire; SES = socioeconomic status; SPM = Standard
Progressive Matrices; T1 = first trimester of pregnancy; T2 = second trimester of pregnancy; T3 = third trimester of pregnancy; TSCD = Tohoku Study of Child Development; WISC =
Weschler Intelligence Scale for Children; WRAML = Wide Range Assessment of Memory and Learning; WRAT = Wide Range Achievement Test; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
tStudies published since the 2013 Integrated Science Assessment for Lead.
Table 3-4T Animal toxicological studies of Pb exposure and cognitive function
Studv Species (Stock/Strain), n, Timing of
" Sex Exposure
Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
Corv-Slechta et al. (2012) Rat (Lonq-Evans) GD-60 to 10 mo
Control (tap water), M, n =
12
Oral, drinking water
Oral, lactation
In utero
PND 5-6:
<5 pg/dL for Control
12.5 pg/dL for 50 ppm
2-3 mo to 10 mo:
Operant Behavior
50 ppm, M, n = 12
2.5 mo:
<5 pg/dL for Control
6.43 pg/dL for 50 ppm
10 mo:
<5 |jg/dL for Control
8.98 |jg/dL for 50 ppm
3-293
-------
Study Species (Stock/Strain), n, Timingof Exposure Details BLL as Reported (Mg/dL) eSKKS
Zou et al. (2015) Mouse (ICR) ~5wkto8wk Oral, drinking water 8wk: 8 wk: Morris
Control (distilled water), M, water maze
n = 10 1.8 |jg/dL for Control
250 mg/L solution, M, n = 21.7 |jg/dL for 250 mg/L
10
3-294
-------
Study Species (Stock/Strain), n, Timingof Exposure Details BLL as Reported (Mg/dL) eSKKS
Corv-Slechta et al. (2013) Mouse (C57BL/6) GD-60 to 12 mo
Oral, drinking water
PND 75 - Females:
7-12 mo:
Control (distilled deionized
Oral, lactation
Operant Behavior
water) - NS, M/F, n = 10-
In utero
-------
Study Species (Stock/Strain), n, Timingof Exposure Details BLL as Reported (Mg/dL) eSKKS
Weston etal. (2014)
Rat (Long-Evans) GD -60 to PND 21
Control (tap water), M/F, n
= 22(11/11)
50 ppm, M/F, n = 22 (11/11)
Oral, lactation
In utero
PND 5-6 - Males:
0.76 pg/dL for Control
15.7 pg/dL for 50 ppm
PND 5-6 - Females:
0.82 pg/dL for Control
14.7 pg/dL for 50 ppm
>PND 60:
Operant Behavior
Betharia and Maher
Rat (Sprague Dawley) GD 0 to PND 20
Oral, lactation
PND 2:
PND 21-25, 56-
(2012)
PND 21-25:
In utero
60: Morris water
1.77 ng/g (0.188 pg/dL) for
maze
Control (RO Dl water), M/F,
Control
n = 11-13
85.17 ng/g (9.02 pg/dL) for 10
10 pg/mL, M/F, n = 11-13
pg/mL
PND 56-60:
PND 25:
Control (RO Dl water), M/F,
0.83 ng/g (0.088 pg/dL) for
n = 9-11
Control
10 pg/mL, M/F, n = 9-11
9.21 ng/g (0.98 pg/dL) for 10
pg/mL
PND 60:
0.23 ng/g (0.024 pg/dL) for
Control
0.30 ng/g (0.032 pg/dL) for 10
pg/mL
3-296
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
Han et al. (2014)
Rat (Wistar)
Control (tap water), M, n =
2 mM - PW, M, n = 8
2 mM - ME, M, n = 8
PW group: PND 21
8 to PND 42
ME group: GD -21
to PND 20
Oral, drinking water
Oral, lactation
In utero
PND 21:
7.36 |jg/L (0.74 pg/dL) for Control
NR for 2 mM - PW
146.6 pg/L (14.7 pg/dL) for 2 mM
-ME
PND 63:
9.22 pg/L (0.92 pg/dL) for Control
147.9 pg/L (14.8 pg/dL) for 2 mM
-PW
46.13 pg/L (4.6 pg/dL) for 2 mM -
ME
PND 63 to PND
68: Morris water
maze
Flores-Montova et al.
(2015)
Mouse (C57BL/6) GD 0 to PND 28
Control (Sodium treated
water), M/F, n = 10 (8/2)
30 ppm, M/F, n = 10 (5/5)
330 ppm, M/F, n = 13 (7/6)
Oral, drinking water
PND 28 - Females:
0.02 pg/dL for Control
2.63 pg/dL for 30 ppm
12.92 pg/dL for 330 ppm
PND 28-Males:
0.31 pg/dL for Control
3.10 pg/dL for 30 ppm
15.21 pg/dL for 330 ppm
PND 28: Novel
Odor Recognition
Rahman et al. (2012a)
Rat (Wistar)
Control, M/F, n = 6
0.2% solution (0.002 g/mL),
M/F, n = 10
PND 1 to PND 21 Oral, drinking water
PND 21:
1.35 pg/dL for Control
12.40 pg/dL for 0.2% solution
PND 21: Morris
water maze
3-297
-------
Study
Species (Stock/Strain), n,
Sex
Exposure Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
Rahman et al. (2012b)
Rat (Wistar)
PND 1 to PND 30 Oral, drinking water
PND 21:
PND 21, 30:
Control (tap water), M/F, n
Oral, lactation
Morris water
= 6
1.4 pg/dL for Control
maze
0.2% solution, M/F, n = 10
12.1 pg/dL for 0.2% solution
PND 30:
1.2 pg/dL for Control
12.8 pg/dL for 0.2% solution
Mansouri et al. (2012)
Rat (Wistar)
Control (distilled water),
M/F, n = 16 (8/8)
50 mg/L, M/F, n = 16 (8/8)
PND 70 to PND 100 Oral, drinking water
PND 100-Males:
2.05 pg/dL for Control
8.8 pg/dL for 50 mg/L
PND 100 - Females:
2.17 pg/dL for Control
6.8 pg/dL for 50 mg/L
PND 100: Morris
water maze,
Novel Object
Recognition
3-298
-------
Anderson et al. (2016)
Rat (Long-Evans)
Control (untreated), M/F, n
= 16 (8/8)
150 ppm, M/F, n = 16 (8/8)
per duration
375 ppm, M/F, n = 16 (8/8)
per duration
750 ppm, M/F, n = 16 (8/8)
per duration
Perinatal exposure
group: GD -10 to
PND 21
Early postnatal
exposure group:
PND 0 to PND 21
Long-term postnatal
exposure group:
PND 0 to PND 55
Oral, diet
Oral, lactation
In utero
3-299
PND 65 - Perinatal exposure
females:
0 |jg/dL for Control
1.36 |jg/dL for 150 ppm
2.13 |jg/dL for 375 ppm
2.08 |jg/dL for 750 ppm
PND 65 - Early postnatal
exposure females:
0 |jg/dL for Control
2.11 |jg/dL for 150 ppm
2.0 |jg/dL for 375 ppm
3.09 |jg/dL for 750 ppm
PND 65 - Long-term exposure
females:
0 |jg/dL for Control
4.5 |jg/dL for 150 ppm
5.75 |jg/dL for 375 ppm
9.58 |jg/dL for 750 ppm
PND 65 - Perinatal exposure
males:
0 |jg/dL for Control
1.25 |jg/dL for 150 ppm
2.42 |jg/dL for 375 ppm
2.47 |jg/dL for 750 ppm
PND 55, 56, 57,
and 65: Trace
Fear Conditioning
-------
Study Species (Stock/Strain), n, Timingof Exposure Details BLL as Reported (Mg/dL) eSKKS
PND 65 - Early postnatal
exposure males:
0 |jg/dL for Control
1.64 |jg/dL for 150 ppm
1.95 |jg/dL for 375 ppm
2.83 |jg/dL for 750 ppm
PND 65 - Long-term exposure
males:
0 |jg/dL for Control
2.01 |jg/dL for 150 ppm
8.0 |jg/dL for 375 ppm
7.46 |jg/dL for 750 ppm
7.61 |jg/L (0.76 |jg/dL) for Control PND 35^0:
Morris water
84.3 |jg/L (8.43 |jg/dL) for 300 maze
ppm
Mena et al. (2016) Rat (Sprague Dawley) PND 0 to PND 21 Oral, lactation
Control (deionized water),
M/F, n = not specified
300 ppm Pb, M/F, n = not
specified
3-300
-------
Study
Species (Stock/Strain), n, Timing of
Sex Exposure
Exposure Details BLL as Reported (pg/dL)
Li etal. (2016a)
Mouse (Kunming)
GD 1 to PND 21 Oral, drinking water
PND 21:
PND 21, 22, 23,
Control (untreated), M/F, n
24, 25, 26: Morris
= 10
9.8 |jg/L (0.98 pg/dL) for Control
water maze
0.1% solution (1000 ppm),
42.5 |jg/L (4.25 pg/dL) for 1,000
M/F, n = 10
ppm
0.2% solution (2000 ppm),
85.3 pg/L (8.53 pg/dL) for 2000
M/F, n = 10
ppm
0.5% solution (5000 ppm),
106.4 pg/L (10.64 pg/dL) for 5000
M/F, n = 10
ppm
Li etal. (2016c)
Mouse (Kunming)
GD 0 to PND 21 Oral, lactation
PND 21:
PND 21: Morris
Control (distilled water),
In utero
water maze
M/F, n = 10
10.62 pg/L (1.1 pg/dL) for Control
0.1% solution (mass
40.71 pg/L (4.1 pg/dL) for 0.1%
fraction), M/F. n = 10
solution
0.2% solution (mass
81.77 pg/L (8.2 pg/dL) for 0.2%
fraction), M/F, n = 10
solution
0.5% solution (mass
103.36 pg/L (10.3 pg/dL) for
fraction), M/F, n = 10
0.5% solution
Mena et al. (2016)
Rat (Sprague Dawley)
PND 1 to PND 21 Oral, lactation
PND 35:
NR: Morris water
Control (deionized water),
maze
M/F, n = 7
5.6 pg/L (0.56 pg/dL) for Control
300 ppm, M/F, n = 7
84.84 pg/L (8.48 pg/dL) for 300
ppm
3-301
-------
Study Species (Stock/Strain), n, Timingof Exposure Details BLL as Reported (Mg/dL) eSKKS
Wana et al.
(2013)
Rat (Sprague Dawley)
GD Oto PND 1,
Oral, drinking water
PND 72:
PND 65 to PND
Control (untreated), M/F, n
PND 1 to PND 21,
Oral, lactation
69: Morris water
= 6
PND 21 to 42
In utero
34.99 |jg/L (3.5 pg/dL) for Control
maze
0.2% solution (w/v), M/F, n
35.78 pg/L (3.58 pg/dL) for 0.2 %
= 6 - Gestational Exposure
solution Gestational
0.2% solution (w/v), M/F, n
65.97 pg/L (6.60 pg/dL) for 0.2%
= 6 - Lactational Exposure
solution Lactational
0.2% solution (w/v), M/F, n
110.67 pg/L (11.07 pg/dL) for
= 6 - Ablactational
0.2% solution Ablactational
Exposure
Wana et al.
(2016)
Rat (Sprague Dawley)
PND 24 to PND 56
Oral, drinking water
PND 56:
PND 60-66:
Control (tap water), M, n = 7
Trace Fear
11 pg/L (1.1 pg/dL) for Control
Conditioning
100 ppm, M, n = 9
133 pg/L (13.3 pg/dL) for 100
ppm
Zhanq et al.
(2014)
Mouse (Kunming)
GD Oto PND 21
Oral, lactation
PND 36:
PND 29, 30:
Control (distilled water),
In utero
Passive
M/F, n = 12
18.5 pg/L (1.9 pg/dL) for Control
Avoidance Test,
PND 31-35:
0.4% solution, M/F, n = 13
136.7 pg/L (13.7 pg/dL) for 0.4%
Morris water
Solution
maze
3-302
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
Barkur and Bairv (2015b) Rat (Wistar) GD -30 to PND 21 Oral, lactation
Control (untreated), M, n = In utero
6
0.2% solution - Gestational,
M, n = 6
0.2% solution - Lactational,
M, n = 6
0.2% solution - Gestation +
Lactation, M, n = 6
0.2% solution -
Pregestational, M, n = 6
PND 22:
0.18 pg/dL for Control
3.02 pg/dL for 0.2% solution -
Pregestation
5.30 pg/dL for 0.2% solution -
Gestational
26.65 pg/dL for 0.2% solution -
Lactational
32.0 pg/dL for 0.2% solution -
Gestation + Lactation
PND 30 to PND
36: Morris water
maze, PND 26,
27, 28: Passive
Avoidance Test
Barkur et al. (2011)
Rat (Wistar) GD 0 to PND 21
Control (tap water), M, n = 9
0.2% solution (w/v), M, n =
9
Oral, lactation
In utero
PND 120:
0.24 pg/dL for control
0.47 pg/dL for 0.2% solution
PND 120:
Passive
Avoidance Test
3-303
-------
Verma and Schneider Rat (Long-Evans)
(2017) Control, M/F, n = 32 (16/16)
PERI: GD -14 to Oral, lactation
PND 21 In utero
150 ppm chow (PERI), M/F,
n = 32 (16/16)
150 ppm chow (EPN), M/F,
n = 32 (16/16)
EPN: PND 0 to PND
21
3-304
PND 14-PERI Males:
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
-------
Verma and Schneider Rat (Sprague Dawley)
(2017) Control, M/F, n = 36 (18/18)
PERI: GD -14 to Oral, lactation
PND 21 In utero
150 ppm Chow (PERI),
M/F, n = 36 (18/18)
150 ppm Chow(EPN), M/F,
n = 36 (18/18)
EPN: PND 0 to PND
21
3-306
PND 14-PERI Males:
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
-------
Anderson et al. (2012) Rat (Long-Evans) GD-10toPND21 Oral, lactation
Control (untreated chow),
M/F, n = 28 (11/17)
250 ppm Chow, M/F, n = 15
(10/5)
750 ppm Chow, M/F, n = 25
(13/12)
1500 ppm Chow, M/F, n =
23 (12/11)
PND 1 - Males: PND 55: Morris
water maze
0 |jg/dL for Control
18.9 |jg/dLfor250 ppm
52.5 |jg/dL for 750 ppm
52.5 |jg/dL for 1500 ppm
PND 1 - Females:
0 |jg/dL for Control
21.9 |jg/dLfor250 ppm
47.2 |jg/dL for 750 ppm
56.7 |jg/dL for 1500 ppm
PND 7-Males:
0 |jg/dL for Control
8.5 |jg/dL for 250 ppm
29.1 |jg/dL for 750 ppm
35.7 |jg/dL for 1500 ppm
PND 7 - Females:
0 |jg/dL for Control
14.7 |jg/dLfor250 ppm
26.9 |jg/dL for 750 ppm
37.6 |jg/dL for 1500 ppm
PND 14- Males:
0 |jg/dL for Control
3-308
-------
Study Species (Stock/Strain), n, Timingof Exposure Details BLL as Reported (Mg/dL) eSKKS
10.5 [jg/dL for 250 ppm
18.6 [jg/dL for 750 ppm
24.8 [jg/dL for 1500 ppm
PND 14 - Females:
0 [jg/dL for Control
11.8 [jg/dL for 250 ppm
20.2 [jg/dL for 750 ppm
26.4 [jg/dL for 1500 ppm
PND 21 - Males:
0 [jg/dL for Control
18.6 [jg/dL for 250 ppm
28.8 [jg/dL for 750 ppm
28.7 [jg/dL for 1500 ppm
PND 21 - Females:
0 [jg/dL for Control
17.9 [jg/dL for 250 ppm
27.4 [jg/dL for 750 ppm
29.8 [jg/dL for 1500 ppm
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Zhao et al. (2018) Rat (Sprague Dawley) GD-14toPND10 Oral, lactation
Control (tap water), M, n = 8 In utero
0.005% solution, M, n = 8
0.01% solution, M, n = 8
0.02% solution, M, n = 8
3-310
PND 0:
1.9 |jg/dL for Control
17.9 |jg/dL for 0.005% solution
23.2 |jg/dL for 0.01% solution
48.8 |jg/dL for 0.02% solution
PND 3:
I.9 |jg/dL for Control
6.7 |jg/dL for 0.005% solution
II.5 |jg/dL for 0.01% solution
23.1 |jg/dL for 0.02% solution
PND 7:
1.3 |jg/dL for Control
8.1 |jg/dL for 0.005% solution
12.3 |jg/dL for 0.01 % solution
18.7 |jg/dL for 0.02% solution
PND 10:
1.2 |jg/dL for Control
5.6 |jg/dL for 0.005% solution
7.0 |jg/dL for 0.01% solution
12.3 |jg/dL for 0.02% solution
PND 14:
0.7 |jg/dL for Control
PND 30: Morris
water maze
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
4.0 pg/dL for 0.005% solution
5.5 pg/dL for 0.01% solution
8.9 pg/dL for 0.02% solution
PND 21:
1.1 pg/dL for Control
2.5 pg/dL for 0.005% solution
2.5 pg/dL for 0.01% solution
2.98 pg/dL for 0.02% solution
PND 30:
1.5 pg/dL for Control
1.0 pg/dL for 0.005% solution
1.5 pg/dL for 0.01% solution
1.5 pg/dL for 0.02% solution
Neuwirth et al. (2019b)
Rat (Long-Evans)
Control (tap water), M/F, n
= 12 (6/6)
363.83 |jM solution, M/F, n
= 12 (6/6)
GD Oto PND 22
Oral, lactation
In utero
PND 22:
NR for Control
5.3-15 pg/dL for 364 pM solution
PND 56-90:
ND for Control,
ND for 364 pM
PND 56-90:
Attention Set
Shifting Test
3-311
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Study Species (Stock/Strain), n, Timingof Exposure Details BLL as Reported (Mg/dL) eSKKS
Neuwirth et al. (2019c)
Rat (Long-Evans)
PERI: GD -14 to
Oral, lactation
PND 14 - Females:
NR: Attention Set
Control, M/F, n = 12 (6/6)
PND 22
In utero
Shifting Test
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
Tartaqlione et al. (2020) Rat (Wistar)
GD -28 to PND 23
Control (tap water), M/F n =
16 (9/7)
50 mg/L, M/F, n = 16 (9/7)
Oral, lactation
In utero
PND 23:
0.007 |jg/mL (0.7 pg/dL) for
Control
0.255 pg/mL (25.5 pg/dL) for 50
mg/L
PND 35: Y Maze
- Spontaneous
Alternation, PND
63-65: Novel
Object
Recognition, PND
68-72: Morris
water maze
Xiao et al. (2014)
Rat (Wistar)
Control (tap water), M/F, n
= 10 (5/5)
Pre-weaning: 2 mM
solution, M/F, n = 10 (5/5)
Postweaning: 2 mM
solution, M/F, n = 10 (5/5)
Pre-weaning: GD
-21 to PND 21
Postweaning: PND
21 to PND 84
Oral, drinking water
Oral, lactation
In utero
PND 21 - Pre-weaning:
10.09 pg/L (1 pg/dL) for Control
103.8 pg/L (10.4 pg/dL) for 2 mM
solution
PND 21 - Postweaning:
Not Reported
PND 91 - Pre-weaning:
10.32 pg/L (1 pg/dL) for Control
39.27 pg/L (3.9 pg/dL) for 2 mM
solution
PND 91 - Postweaning:
10.32 pg/L (1 pg/dL) for Control
105.45 pg/L (10.5 pg/dL) for 2
mM solution
PND 85 to 90:
Morris water
maze
3-313
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
Sobolewski et al. (2020) Mouse (C57BL/6)
F0:
Control (distilled Dl water),
F, n = 10
100 ppm, F, n = 10
F1:
see Figure 1, n = 12
F2:
see Figure 1, n = 12
F3:
see Figure 1, n = 8-10
F1: GD -60 to PND Oral, lactation
23-27
In utero
F1 PND 6-7:
0 pg/dL for Control
12.5 pg/dL for 100 ppm (F0
dosing)
F3 PND 6-7:
0 ng/dL for Control
0 pg/dL for 100 ppm (F0 dosing)
PND 60-120
(variable by
endpoint): Fl
Training
Ouvana et al. (2019)
Rat (Sprague Dawley) GD 0 to PND 679
Control (tap water), M/F, n
= 6-10
0.05/0.01% solution, M/F, n
= 6-10
Oral, drinking water
Oral, lactation
In utero
wk 97:
0 mg/L (0 pg/dL) for Control
0.216 mg/L (21.6 pg/dL) for
0.05/0.01% solution
PND 674 to PND
679: Morris water
maze
Singh et al. (2019)
Rat (Wistar) 3 mo to 6 mo
Control (distilled water), M,
n = 5
2.5 mg/kg, M, n = 5
Oral, gavage
6 mo:
5.76 pg/dL for Control
28.4 pg/dL for 2.5 mg/kg
6 mo: Morris
water maze
Xiao et al. (2020)
Rat (Sprague Dawley)
Control (tap water), F, n
10
125 ppm, F, n = 10
GD -7 to PND 68
Oral, drinking water
Oral, lactation
In utero
PND 68:
24.23 ng/mL (2.4 pg/dL) for
Control
205 ng/mL (20.5 pg/dL) for 125
ppm
PND 56 -61:
Morris water
maze, PND 55: Y
Maze -
Spontaneous
Alternation
3-314
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Study Species (Stock/Strain), n, Timingof Exposure Details BLL as Reported (Mg/dL) eSKKS
Su etal. (2016)
Rat (Sprague Dawley)
Control (deionized water
with 0.9% saline), M, n = 15
200 ppm, M, n = 16
PND20toPND76 Oral, gavage
PND 76:
7.99 |jg/L (0.8 |jg/dL) for Control
84.17 |jg/L (8.4 pg/dL) for 200
ppm
PND 76: Morris
water maze
3-315
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An et al. (2014) Rat (Sprague Dawley) 4 wk to 12 wk
Control (deionized water
with NaAc), M, n = 12
100 ppm, M, n = 12
200 ppm, M, n = 12
300 ppm, M, n = 12
Oral, drinking water
5wk: 12-wk: Morris
water maze
0.96 |jg/dL for Control
7.07 |jg/dL for 100 ppm
11.54 |jg/dLfor200 ppm
14.76 |jg/dL for 300 ppm
6 wk:
0.96 |jg/dL for Control
8.13 |jg/dL for 100 ppm
12.92 |jg/dL for 200 ppm
16.65 |jg/dL for 300 ppm
7 wk:
0.96 |jg/dL for Control
9.68 |jg/dL for 100 ppm
13.37 |jg/dL for 200 ppm
19.48 |jg/dL for 300 ppm
8 wk:
0.96 |jg/dL for Control
9.64 |jg/dL for 100 ppm
17.07 |jg/dL for 200 ppm
22.02 |jg/dL for 300 ppm
9 wk:
0.96 |jg/dL for Control
3-316
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Study Species (Stock/Strain), n, Timingof Exposure Details BLL as Reported (Mg/dL) eSKKS
12.12 [jg/dL for 100 ppm
20.7 [jg/dL for 200 ppm
22.28 [jg/dL for 300 ppm
10 wk:
0.96 [jg/dL for Control
11.48 [jg/dL for 100 ppm
17.75 [jg/dL for 200 ppm
24.69 [jg/dL for 300 ppm
11 wk:
0.96 [jg/dL for Control
11.51 [jg/dL for 100 ppm
17.52 [jg/dL for 200 ppm
22.18 [jg/dL for 300 ppm
12 wk:
0.96 [jg/dL for Control
11.41 [jg/dL for 100 ppm
17.23 [jg/dL for 200 ppm
22.57 [jg/dL for 300 ppm
3-317
-------
Study
Species (Stock/Strain), n, Timing of
Sex Exposure
Exposure Details BLL as Reported (pg/dL)
Li etal. (2013)
Rat (Wistar)
Control (tap water), M/F, n
= 16 (8/8)
500 ppm, M/F, n = 16 (8/8)
4 wk to 16 wk
Oral, drinking water
4 mo:
29.99 |jg/L (3 pg/dL) for Control
159.54 |jg/L (16 |jg/dL) for 500
ppm
4 mo: Morris
water maze
Zhu et al. (2019b)
Rat (Sprague Dawley)
GD 0 to 12 mo
Oral, drinking water
12 mo:
NR: Morris water
Control (deionized water),
Oral, lactation
maze
M/F, n = 32
In utero
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
Zhang et al. (2012)
Rat (Sprague Dawley)
Control (deionized water),
NR (40-60 g)
Oral, drinking water
+8 wk from start of exposure: +8 wk from start
of exposure:
M, n = 10
49.9 ng/mL (5 pg/dL) for Control
100 ppm, M,
n = 10
100.9 ng/mL (10.1 pg/dL) for 100
ppm
200 ppm, M,
n = 10
128.6 ng/mL (12.9 pg/dL) for 200
300 ppm, M,
n = 10
ppm
147.7 ng/mL (14.8 pg/dL) for 300
ppm
maze
Hong et al. (2021)
Rat (Sprague Dawley) GD 0 to PND 21
Control (tap water), M/F, n
= 50
1 g/L Pb solution, M/F, n =
50
Oral, lactation
In utero
0.009 mg/L for Control, 0.291
mg/L for 1 g/L Pb - PND 21
PND 21-27:
Morris water
maze
Biioor et al. (2012)
Rat (Wistar) GD 0 to PND 45
Control (deionized water),
M/F, n = 10
50 ppm, M/F, n = 10
Oral, drinking water
Oral, lactation
In utero
PND 45:
4.06 pg/dL for Control
10.65 pg/dL for 50 ppm
PND 45: Passive
Avoidance Test
Wang et al. (2021a)
Rat (Sprague Dawley)
Control (deionized water),
M, n = 8
0.05% solution, M, n = 8
0.1% solution, M, n = 8
GD Oto PND 21
Oral, lactation
In utero
PND 21:
23.1 |jg/L (2.31 pg/dL) for Control
248 pg/L (24.8 pg/dL) for 0.05%
solution
302 pg/L (30.2 pg/dL) for 0.1%
solution
PND 21: Morris
water maze
361 pg/L (36.1 pg/dL) for 0.2%
solution
3-319
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Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
BLL as Reported (pg/dL)
Endpoints
Examined
Liu et al. (2022c)
Rat (Sprague Dawley)
Control (tap water), M, n =
10
0.2% solution, M, n = 10
PND 35 to PND 119 Oral, drinking water
PND 119:
10.9 |jg/L (1.09 pg/dL) for Control
176 pg/L (17.6 pg/dL) for 0.2%
solution
PND 119: Morris
water maze
Wang et al. (2021b)
Rat (Sprague Dawley)
Control (deionized water),
M/F, n = 12
GD -28 to PND 21
Oral, lactation
In utero
PND 21:
PND 21: Morris
water maze
23.9 pg/L (2.39 pg/dL) for Control
0.05% solution, M/F, n = 10
206 pg/L (20.6 pg/dL) for 0.05%
solution
Al-Qahtani et al. (2022)
Mouse (Albino)
Control (distilled water), M,
n = 10
0.2 mg/kg, M, n = 10
8-9 wk to 14-15 wk Oral, gavage
14-15 wk:
1.2 pg/100 mL (1.2 pg/dL) for
Control
7.1 pg/100 mL (7.1 pg/dL) for 0.2
mg/kg
NR: Active
Avoidance Test
Long et al. (2022)
Rat (Sprague Dawley)
Control (untreated), M, n
12
200 mg/L solution, M, n =
12
6 wk to 18 wk
Oral, drinking water
18 wk:
2.14 pg/L (0.214 pg/dL) for
Control
32.48 pg/L (3.25 pg/dL) for 200
mg/L solution
NR: Morris water
maze, NR: Active
Avoidance Test
BLL = blood lead level; CI = confidence interval; EPN = early postnatal; F = female; F1 = first filial generation; Fl
male; ME = maternal exposure; mo = month(s); NaAc = sodium acetate; NR = not reported; NS = no stress; Pb =
PW = postweaning; RO Dl = reverse osmosis deionized; SD = standard deviation; wk = week(s).
: fixed interval; GD = gestational day; LOD = limit of detection; M =
lead; PERI = perinatal; PND = postnatal day; PS = prenatal stress;
3-320
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Table 3-5E
Epidemiologic studies of Pb exposure, academic performance, and achievement
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Chandramouli et al.
(2009)
Avon
U.K.
Jul.-Dec. 1992 (birth)
Followed 8 yr
10% random subsample
of Avon Longitudinal
Study of Parents and
Children (ALSPAC)
n = 488
School children
Blood
Earlier childhood venous
blood; AAS using micro
sampling flame
atomization
Age at measurement: 30
mo
Academic achievement
Standardized
Achievement Test
Age at outcome:
7 yr
Maternal education and
smoking, home
ownership, home
facilities score, family
adversity index, paternal
SES, parenting attitudes
at 6 mo, child sex. Also
considered child IQ
Per doubling BLLb
-0.3 (-0.5, -0.1)
Cohort
Mean (SD): NR
Group 1: 0-<2 |jg/dL
Group 2: 2-<5 |jg/dL
Group 3: 5—<10 |jg/dL
Group 4: >10 pg/dL
Miranda et al. (2009)
School children, n=
Blood
Academic achievement
Sex, age of blood Pb
Score vs. blood Pb
57,568
Surveillance database
measurement, race,
category 1 |jg/dL
North Carolina
4th grade EOG test score
enrollment in
free/reduced lunch
2 |jg/dL:
U.S.
Screened for Pb at age
Age at measurement: 9-
for reading (2001-2005)
program, parental
-0.30 (-0.58, -0.01)
9-36 mo in 100 NC
36 mo
education, charter
3 |jg/dL:
1995 through 1999
counties
school.
-0.46 (-0.73, -0.19)
(screening)
4 |jg/dL:
-0.52 (-0.79, -0.24)
Cohort
5 |jg/dL:
-0.80 (-1.08, -0.51)
3-321
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Min et al. (2009)
Cleveland, OH
1994-1996 (birth)
Followed to age 11 yr
Cohort
Birth cohort, n = 267
86% African-American
with high prevalence of
prenatal drug and alcohol
exposure
Blood
Earlier childhood (age 4
yr)
Mean (SD): 7.0 (4.1)
Interval analyzed: 3.0
(10th percentile)-10
WJTA (math and reading Sex, caregiver education, Math score: -2.5 (-4.6,
scores
Age at outcome:
11 yr
family income,
race/ethnicity, test
language.
-0.38)
Reading score: -2.9
(-4.4, -1.4)
Lanphear et al. (2000)
United States
1988-1994
Cross-sectional
U.S. NHANES
n = 4,853 children ages
6-16 yr (born 1972-
1988)
Large U.S.
representative study of
multiple risk factors and
outcomes
Blood
Concurrent
GM (SD): 1.9 (7.0)
63.5% <2.5
Detection limit = 0.5
Interval analyzed: 1-5
WRAT (arithmetic and
reading scores)
Age at outcome: 6-16 yr
Linear regression
Child sex, race/ethnicity,
poverty index ratio,
reference adult
education, serum ferritin
and cotinine levels. Did
not consider potential
confounding by parental
cognitive function or
caregiving quality.
-0.05 (-0.09, -0.01)
Krieq et al. (2010)
United States
1991-1994
Cross-sectional
U.S. NHANES
n = 773 children ages
12-16 yr (born 1972-
1982)
Large U.S.
representative study of
multiple risk factors and
outcomes
Blood
Concurrent
GM (SD): 1.9 (7.0)
63.5% <2.5
Detection limit = 0.5
Interval analyzed: 1-5
WRAT (arithmetic and
reading scores)
Age at outcome: 6-16 yr
Log-linear regression
Child sex, caregiver
education, family income,
race/ethnicity, test
language. Did not
consider potential
confounding by parental
cognitive function or
caregiving quality.
-0.34 (-0.59, -0.08)
3-322
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Surkan et al. (2007)
Boston, Massachusetts
and Farmington, Maine
United States
Cross-sectional
n = 389 children
6-10 yr
Recruitment from trial of
amalgam fillings
Blood
Concurrent
Group 1:1-2
Group 2: 3-4
Group 3: 5-10
Mean (SD): 2.2 (1.6)
WIAT (reading and math
composites)
Age at outcome: 6-10 yr
Caregiver IQ, child age,
SES, race, birth weight.
Also considered potential
confounding by site, sex,
birth order, caregiver
education and marital
status, parenting stress,
and maternal utilization
of prenatal and annual
health care but not
parental caregiving
quality.
-0.69
-6.7 (-
-4.4, 3.0)
12, -1.2)
Chiodo et al. (2007)
Detroit, Michigan
United States
Cross-sectional
n = 495 children (born
1989-1991) age 7 yr
Blood
Concurrent
Mean (SD): 5.0 (3.0)
Test of Early Reading
Ability—2
MAT (math and reading
scores)
Age 7 yr
Maternal
psychopathology, IQ,
prenatal smoking,
prenatal marijuana, SES,
HOME score, caretaker
education and marital
status, # children in
home, child sex. Also
considered child age,
maternal age, custody,
cocaine use, prenatal
alcohol use.
-0.19 (-0.30, —0.08)c
Ferausson et al. (1997) n = 881 children
Christchurch
New Zealand
1977 (birth)
Followed to age 18
Cohort
Tooth Pb (age 6-8 yr)
Mean (SD): 6.2 (3.7) pg/g
Percent leaving school
without school certificate
Ages 16-18 yr
Christchurch Health and
Development Study birth
cohort
Maternal age,
punitiveness, standard of
living, breastfeeding
duration, parental
conflict, grade, residence
on busy roads. Also
considered potential
confounding by sex,
ethnicity, maternal
education, family size,
HOME, SES, ethnicity,
parental change, birth
order, single parent.
0-2 pg/g: 15.6
3-5 pg/g: 16.7
6-8 pg/g: 18.1
9-11 pg/g: 19.7
12+ pg/g: 24.1
p < 0.05
3-323
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Needleman et al. (1990) n = 132 children (1st/2nd Tooth Pb (1st/2nd grade) Failure to graduate high Maternal age at birth,
Failure to graduate: 7.4
Chelsea and Somerville,
MA
United States
1975-1978 (enrollment)
followed to age 18 yr
Cohort
grade) in Massachusetts
schools
distribution
<10 ppm: 50%
10-19.9 ppm: 22.7%
>20 ppm: 27.3%
school
Highest grade achieved
Logistic regression
education, and IQ, family (1.4, 41) d
size, SES, sex, age at
testing, birth order,
alcohol use, mother and
child left hospital
together. Did not
examine potential
confounding by parental
caregiving quality.
OR >20 ppm vs. <10
ppm
Highest grade achieved:
-0.03 (-0.05, -0.01)
per natural log increase
in tooth Pb
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Referencejmd Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tZhana et al. (2013)
Detroit, Michigan
United States
1990-2008 (born)
Followed through grade
3-8 (2008-2010)
Cohort
Students in public
schools in Detroit
n: 21281 (8831-3rd
grade, 7708- 5th grade,
4742- 8th grade)
At least 1 of the 3 tests
(math, science and
reading) taken and BLL
before 6 yr of age
Blood
Venous blood collected
for surveillance by Detroit
Department of Health
and Wellness Promotion
Age at measurement:
Before 6 yr of age (mean
age: 3.1)
Max: Highest BLL before
age 6 yr: 7.12 |jg/dL
Academic achievement
(math, science and
reading) in grade 3, 5,
and 8
Educational attainment in
math, science and
reading on MEAP.
Age at outcome:
3, 5, and 8 grades
Grade level, gender,
race, language, maternal
education, SES (i.e.,
school lunch status).
ORs of Scoring "Less
Than Proficient" on
MEAP Tests (Ref= <1
fjg/dL)
1-5 fjg/dL
Mathematics: 1.42 (1.24,
1.63)
Science: 1.33 (1.10,
1.62)
Reading: 1.45 (1.27,
1.67)
6-10 fjg/dL
Mathematics: 2.00 (1.74,
2.30)
Science: 2.22 (1.82,
2.72)
Reading: 2.21 (1.92,
2.55)
>10 fjg/dL
Mathematics: 2.40 (2.07,
2.77)
Science: 2.26 (1.84,
2.78)
Reading: 2.69 (2.31,
3.12)
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tEvensetal. (2015)
Chicago metropolitan
area, 6 counties
U.S.
1994-1998 (born)
Followed 9-10 yr, 2003-
2006
Cohort
Chicago public school
children
n: 46796
Blood
BLLs obtained from
Chicago Blood Pb
Surveillance program;
ICP-MS or AAS.
Age at measurement:
<72 mo (mean age: 45
mo)
Mean: 4.81 pg/dL
Academic achievement
3rd grade ISAT scores in
Reading and Math; 4
score categories, i.e.,
failure, below standard,
meets standard and
exceeds standard.
Age at outcome:
9-10 yr
Sex, mother's education,
low-income, very low
birth weight/preterm,
child's age at time of
BLL, ISAT vs. Iowa, race
(Interaction with race
ethnicity explored).
RR
Reading Failure
1 iJg/dL increase
All Children:
1.06 (1.05, 1.07)
NH White:
1.14 (1.08, 1.20)
NH Black:
1.05 (1.04, 1.06)
Hispanic:
1.08 (1.05, 1.11)
Math Failure
11Jg/dL increase
All Children:
1.06 (1.05, 1.07)
NH White:
1.11 (1.05, 1.18)
NH Black:
1.05 (1.04, 1.06)
Hispanic:
1.09 (1.06, 1.12)
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Referencejmd Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tBIackowicz et al. (2016)
Chicago
U.S.
1994-1998 (birth)
Followed through 2003-
2006 (3rd grade)
Cohort
School children
n: 12319
Chicago Public Schools.
Blood
Chicago Blood Pb
Registry provided data
on BLL measured
between birth and 2006
Age at measurement:
between birth and 2006
(most recent was used in
analysis)
4.16 |jg/dL
School performance
3rd grade performance
based on ISAT scores
Age at outcome:
3rd grade
Child sex, maternal
education, low-income,
preterm birth, small for
gestational age, child's
age at time of BLL, ISAT
vs. Iowa, and Hispanic
subgroup (Mexican-
American vs. other
Hispanic and Puerto
Rican vs. Other
Hispanic);
Beta
Reading scores: -0.11
(-0.134, -0.086)
Math scores: -0.096
(-0.12, -0.072)
RR
Reading failure: 1.07
(1.05, 1.10)
Math failure: 1.09 (1.06,
1.12)
tShadbeaian et al.
(2019)
North Carolina Statewide
U.S.
1990-2004 (birth)
Followed 6 yr (3-8
grade)
Cohort
NC Pb Poisoning
Prevention Program
Cohort
n: 560,624 (54% of the
Pb surveillance registry)
Living in NC between
2000-2012 with BLL <10
|jg/dL at 0-5 yr
Blood
Child blood (BLL <10,
BLL <5, and a matched
group via CEM with BLL
<5 |jg/dL)
Age at measurement:
0-5 yr
Full sample (BLL <10
|jg/dL) mean: 3.66, Full
sample (BLL <5 |jg/dL)
mean: 2.89, CEM
Matched sample (BLL <5
|jg/dL) mean: 2.40
Academic achievement
Percentile scores on
standardized EOG tests
for math and reading
Age at outcome:
Grade 3 and Grade 8
Child's sex,
race/ethnicity, SES,
Medicaid enrollment,
birth month, and age
upon entry to grade 3,
mother's age, marital
status, parental alcohol
and tobacco use, highest
educational achievement
at the time of the child's
birth, vector representing
school, grade and year
combination
CEM to balance
distributions between
groups
Beta
Decrease in Test-Score
Percentile in Children
with 5-6 |jg/dL vs. BLL <
1 |jg/dL
Math: 0.95 (0.66, 1.24)
Reading: 1.41 (1.12,
1.70)
2- and 3-way interactions
for BLL*grade,
covariates*grade,
BLL*grade*covariates.
3-327
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Referencejmd Study study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tSkerfvina et al. (2015)
Landskrona and
Trelleborg
Sweden
1978-2007 (enrollment
during primary school)
Followed through age 16
Cohort
Primary school children
n: 3176
Blood
Child venous blood;
flame or electrothermal
atomization AAS
Age at measurement:
7-12 yr
34 |jg/L; Median: 30
75th: 44
90th: 60
Max: 162
School performance (see
also Section 3.6.1 [adult
cognitive function])
School performance after
nine-year compulsory
schooling. 4-5 categories
from not passing to
passing with merit (based
on ranking)
Age at outcome:
16 yr
Child and parent country
of birth, parental
education, total family
income, father's IQ.
Beta
Merits
Children (BLL < 5 |jg/dL):
-10.9 (-15.486, -6.314)
All Children: -6.36
(-9.986, -2.734)
Grades
Children (BLL < 5 pg/dL):
-0.112 (-0.177, -0.047)
All Children: -0.155
(-0.21, -0.1)
Note: CIs estimated from
p-values.
AAS = atomic absorption spectrometry; BLL = blood lead level; CEM = coarsened exact matching; CI = confidence interval; EOG = end of grade; GM = geometric mean; HOME =
Health Outcomes and Measures of the Environment; ICP-MS = inductively coupled plasma mass spectrometry; ISAT = Illinois Standard Achievement Test; IQ = intelligence quotient;
MAT = Metropolitan Achievement Test; MEAP = Michigan Educational Assessment Program; Ml = Michigan; NHANES = National Health and Nutrition Examination Survey; NR = not
reported; Pb = lead; SD = standard deviation; SES = socioeconomic status; WIAT = Wechsler Individual Achievement Test; WJTA = Woodcock-Johnson Test of Achievement; WRAT
= Wide Range Achievement Test; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResults are not standardized (e.g., BLL distribution data needed to calculate the standardized estimate was not reported or categorical data was analyzed).
The CI was calculated from a p-value and the true CI may be wider or narrower than calculated.
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-328
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Table 3-6E Epidemiologic studies of Pb exposure and cognitive effects: population or group mean blood Pb
levels >5 ugldL
RefereDCesfgnnd StUdy Study Population Exposure Assessment
Al-Saleh et al. (2020)
Saudi Arabia
2011-2013 (enrollment)
Followed through 2011 —
2013 and 2017-2018
Cohort
Lactating mother-infant
pairs
n: 82 (36 males and 46
female children).
Blood, Hair, Urine, Breast
Milk
Maternal blood, spot
urine, breast milk, and
hair, child spot urine and
hair; AAS with
electrothermal atomizer
Age at measurement:
Maternal measurements
made during lactation
Infants at 3-12 mo
(lactation) and children at
5-8 yr old
Lactation:
GM: maternal urine
:5.881 |jg/L, hair :1.717
|jg/g, blood GM: 2.346
|jg/dL, breastmilk: 46.483
|jg/L; Infant urine: 4.946
|jg/L, hair: 2.894 |jg/g;
Outcome
Neurodevelopmental
performance and visual-
motor integration (Test of
Nonverbal Intelligence
2nd edition and Beery
VMI 3rd edition,
respectively.)
Age at outcome:
5-8 yr old
Confounders
Child's age and sex,
maternal age, BMI,
parity, lifestyle,
educational level, SES,
residential
characteristics, urinary
cotinine levels (an index
of exposure to
secondhand smoke).
Effect Estimates and
95% CIs
Beta (95% Cl)a
BVMI: 0.012 (-1.989,
2.014)
TONI: -0.044 (-1.809,
1.721)
Early childhood:
GM: urine: 2.563 |jg/L,
hair: 0.850 |jg/g max:
urine: 20.826 |jg/L, hair:
4.470 |jg/g
3-329
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Bara et al. (2018)
Montevideo
Uruguay
Study years NR
Cross-sectional
n: 206
Children living in areas
considered high risk for
metal exposure.
Blood
Child venous blood;
flame AAS or GFAAS
Age at measurement:
5-8 yr
4.2 |jg/dL
Executive function
BRIEF (teacher rating)
Age at outcome:
5-8 yr
Child IQ, iron status, and
BMI, blood Pb testing
method, household
possessions, maternal
education, current parent
smoking.
PR (95% Cl)b
BRIEF- Global Executive
Composite
Children with BLL >5 vs.
<5 |jg/dL: 1.02 (0.96,
1.09)
Boys: 1.00 (0.98, 1.01)
Girls: 1.01 (0.99, 1.04)
Cai et al. (2021)
Guangxi
China
Study years NR
Cross-sectional
School children
n: 255
Participants living near a
Pb and zinc mine (-500
m distance between
school and mine).
Blood
Child venous blood (Pb
intoxication and non-Pb
intoxication groups);
GFAAS
Age at measurement:
7-12 yr
Median Pb level: Rice
samples: 0.10 mg/kg,
Blood: 84.8 |jg/L
75th: Blood: 115.4 pg/L
Max: Rice: 0.53 mg/kg
Perception and
reasoning
Raven's SPM
Age at outcome:
7-12 yr
Age, gender, physical
condition, lifestyle habits,
educational attainment
and smoking habit of
parents, family
environment and
economy.
Beta (95% Cl)b
IQ, RSPM: -0.58 (-1.031,
-0.129)
Jeona et al. (2015)
Multi-center
South Korea
Cross-sectional
May 2006-Dec 2010
MOCEH study
n: 194
Birth cohort- mother-
infant pairs followed
through 60 mo of age.
Cross-sectional analysis
conducted.
Blood
Child venous blood;
GFAAS
Age at measurement:
60 mo
GM: 13.01 pg/L
Max: 35.05 pg/L
FSIQ, VIQ, PIQ (Korean
WPPSI-R)
Age at outcome:
60 mo
Sex, parental education,
family income,
breastfeeding status,
CRP level, mother's BLL
during pregnancy
Note: Mediation analysis
to examine the
relationship of BLL, iron
deficiency and IQ.
Beta (95% CI)b
Verbal IQ and In-BLL
(pg/L): -9.587 (-16.829,
-2.344)
3-330
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Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Kao et al. (2021)
Taipei
Taiwan
2011-2014
Cross-Sectional
recruited from Taipei
MacKay Memorial
Hospital
n:139 children less than
3 yr of age
Hair, fingernails
Child hair, fingernails;
ICP-MS
Age at Measurement:
Mean (SD) 2.8 (0.4)
years (children under 3
yr)
Mean (SD): hair 2.9 (4.8)
pg/g,
nails 0.8 (5.1) |jg/g
BSID-III cognitive and
language development
scores
Age at outcome: 2.8 ±
0.4 yr
General linear models
adjusted for sex,
gestational age at birth,
age of the house (years),
leafy-vegetable intake
(servings/week), and the
area of surface roads
within 100 m of the
residence
Regression results were
not reported because
they were not statistically
significant.
3-331
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Kordas et al. (2011)
Mexico City, Mexico
Jan 1994-June 1995
Followed for 48 mo
Cohort
Birth cohort
n: 24 mo = 220, 48 mo =
186
Mother-infant pairs from
3 hospitals serving low-
and middle-income
women
Blood
Maternal, cord blood, and
child blood; GFAAS
Age at measurement:
Delivery (maternal, cord),
24 and 48 mo (child)
Mean: Maternal BLL at
delivery: 8.6 |jg/dL, Cord
blood: 6.6, BLL at 24 mo:
8.1, BLL at 48 mo: 8.1
Neurodevelopment using
BSID-II (MDI), MSCA
(general cognitive index
and memory scale).
Age at outcome:
24 (BSID) and 48 mo
(MSCA)
OLS linear regression
Birth weight, gestational
age, child sex; maternal
age, years of schooling,
IQ, smoking status,
marital status crowding in
the house, type of floor in
the house.
(Stratified analysis by
child development 48 mo
and gene polymorphism
also conducted.)
Beta (95% Cl)b
McCarthy Scales of
Children's Abilities, 48
mo
GCI
Concurrent BLL: -0.6
(-0.992, -0.208)
Cord BLL: -0.2 (-0.788,
0.388)
Memory Score
Concurrent BLL: -0.3
(-0.496, -0.104)
Cord BLL: 0.1 (-0.096,
0.296)
Bayley Scales of Infant
Development II, 24 mo
MDI
Concurrent BLL: -0.1
(-0.492, 0.292)
Cord BLL: -0.7 (-1.288,
-0.112)
Kuana et al. (2020)
Nanjing
China
2012
Cross-sectional
Public primary school
children
n: 742
Excluded students with
congenital mental
retardation (third-degree
relatives included) and
diseases.
Blood
Child venous blood; ICP-
MS
Age at measurement:
7-11 yr
Mean: 30.4 |jg/L; Median:
26.1 |jg/L
School performance
Standardized scores on
Chinese, Math and
English added for total
scores.
Age at outcome:
7-11 yr
Age and sex (tests
administered on the
same day).
Beta (no p-value, Cis,
or SE reported)13
Total: -0.168
Chinese: -0.042
Math: -0.039
English: -0.087
3-332
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Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Lee etal. (2017)
Korea
Enrolled 2006-2015,
followed to age 5 (2015)
Cohort
Mothers' and Children's
Environmental Health
(MOCEH)
n: 251
Maternal Blood
GFAAS with Zeeman
background correction
Age at Measurement:
At birth (cord blood)
GM 0.957 |jg/dL
Max: 3.17 |jg/dL
Cognitive Development
The mental
developmental index
(MDI) of the Korean
BSID-II (K-BSID-II) was
administered to infants
who were 6, 12, 24, and
36 mo-old. The Korean
language version of the
Wechsler Preschool and
Primary Scale of
Intelligence - Revised (K-
WPPSI-R) was
administered to children
at 60 mo.
Partial correlation
analysis adjusted for
maternal education, sex
of child, and family
income.
Strongest correlations
between scores
measured at closest
time. Scores more stable
in those at extreme ends
of cognitive development.
3-333
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Liu etal. (2013a)
Jiangsu province
China
2004-2005 (enrollment)
followed 5 yr
Cohort
China Jintan Child
Cohort Study
n: 1341 children (603
girls and 738 boys)
Community based cohort
of preschool children.
Blood
Early child blood;
GFAAS.
Age at measurement:
3, 4 or 5 yr
mean: 6.43 |jg/dL
FSIQ, VIQ, PIQ (Chinese
version ofWPPSI-R).
See also Table 3-5
(School performance
was assessed by
standardized tests on 3
major subjects: Chinese,
English and Math.)
Age at outcome:
6 yr (IQ), 8-10 yr (school
performance)
Child age at blood Pb
test, child gender,
residence as defined as
school location, blood
iron level, parent
education, parent
occupation, and father's
smoking.
Beta for log-transformed
BLL (95% Cl)b (ref: <8
pg/dL)
FSIQ
8-10 |jg/dL: -1.28
(-4.01, 1.46)
>10 |jg/dL: -1.45 (-3.50,
0.67)
Chinese score
8-10 |jg/dL: -3.20
(-5.78, -0.63)
>10 |jg/dL: -4.02 (-7.11,
-0.93)
Math score
8-10 |jg/dL: -5.25
(-8.14, -2.36)
>10 |jg/dL: -5.27 (-8.73,
-1.81)
English score
8-10 |jg/dL: -4.33
(-7.32, -1.34)
>10 |jg/dL: -5.18 (-8.76,
-1.59)
Liu etal. (2018b)
PROGRESS study
Blood
Cognitive development
SES, mother's NR
n: 665
using BSID-III. BSID
hemoglobin during the
Mexico City, Mexico
Maternal blood; joint
scores were centered
second trimester of
Followed for 24 mo
Mother-infant pairs.
exposure to Mn, Pb, Co,
and scaled and
pregnancy, mother's
Cohort
Cr, Cs, Cu, As, Cd, and
presented as z-scores
educational level, child
Sb
normalized to expected
mean of 100 and SD of
gender, mother's WASI
IQ, and Fenton's birth
Age at measurement:
15.
weight z-scores.
Prenatal exposure (2nd
trimester)
Age at outcome:
NR
6, 12, 18, and 24 mo
3-334
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Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Marques et al. (2014)
State of Rondonia,
Western Amazon
Brazil
Cohort
2007-2012
Population-based cohort
n: 96 (TOKS = 51 and
Itapua = 45)
Low SES populations
living in rural and urban
areas including children
living in the vicinity of
TOKS (i.e., multiple
metal exposure).
Breastmilk
Breastmilk; GFAAS
Age at measurement:
6 mo of breastfeeding
(from Marques 2013c)
TOKS: 10.04 pg/L
(mean), 8.2 pg/L
(median); 29.4 pg/L
(max)
Itapua: 3.89 pg/L (mean)
2.5 pg/L (median), 16.2
pg/L (max)
Neurodevelopment
(milestones including age
of walking and talking,
Bayley MDI and PDI);
milestones assessed
based on mothers'
recollection at the time of
visit.
Age at outcome:
6 and 24 mo (Bayley
MDI, PDI)
Birth weight, income,
maternal education,
breastfeeding status.
Beta (95% Cl)b
MDI 6 M -0.293 (-0.50,
0.08)
MDI 24 M -0.234 (-0.60,
0.13)
PDI 6 M -0.062 (-0.28,
0.16)
PDI 24 M -0.129 (-0.34,
0.08)
Age of walking -0.219
(-0.43, 0.002)
Age of talking -0.066
(-0.28, -0.16)
3-335
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Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Maazamen et al. (2015)
Wisconsin (Milwaukee or
Racine)
United States
Enrollment: born during
1996-2000 (Follow-up of
blood lead level: before
child's third birthday;
Follow-up for WKCE
scores: 4th grade)
Cohort
Wisconsin Children's
Lead Levels and
Educational Outcomes
Project (CLLEO)
n: 1076
Blood
Blood lead records used
to categorize children as
not exposed (<5 |jg/dL)
or exposed (>10 |jg/dL
and <20 |jg/dL)
Age at Measurement:
18-36 mo
43% of sample defined
as exposed
Academic achievement:
Wisconsin Knowledge
and Concepts Exam
(WKCE) math and
reading scores
WKCE reading and math
scores obtained from the
Wisconsin Department of
Public Instruction with
parental consent.
Child gender, race,
parental < HS education,
free lunch program,
English language learner,
and child health rating by
parents (excellent vs.
other); interactions with
Pb tested for each
covariate
Beta for entire
distribution of math
scores'3
OLS = -8.94 (-14.84,
-3.05)
Beta for math scores in
quantilesb
10th percentile = -17.00
(-32.13, -3.27)
50th percentile = -8.00
(-15.24, -0.36)
90th percentile = -4.50
(-10.55, 4.50)
Beta for entire
distribution of reading
scores'3
OLS = -13.66 (-19.94, -
7.37)
Beta for reading scores
in quantilesb
10th percentile = -18.00
(-48.72, -3.32)
50th percentile = -14.50
(-20.72, -5.61)
90th percentile = -7.50
(-15.58, 2.07)
3-336
-------
Rawat et al. (2022) n: 43 Blood
India Blood Pb was measured
Not reported via LeadCare II testing
Cross-Sectional analyser
Age at Measurement:
4-12 yr
GM (SD) 19.93 (9.22)
ug/dL
Max: 37.4 |jg/dL
% change in results from
Draw-A-Person test for
Group B compared to
Group A:
Line characteristic: Thick
and sharp = 41%; Soft =
-32%
Detailing: With = 32%;
Without = -9%
Shading: With = -21%;
Without = 24%
Distortion: With = 50%;
Without = -17%
Colours: Warm = 41%;
Cool = -27%
IQ level, performance on
Draw-A-Person Test
The Draw-A-Person test
and the IQ test were
administered in the study
setting
Age at Outcome:
4-12 yr
There were no
adjustments for
confounders, as simple
statistics were employed.
IQ - Mean (SD) scoreb
Group A (<10 |jg/dL Pb,
n = 9): 122.33 (4.03)
Group B (>10 |jg/dL Pb,
n = 34): 96.03 (12.76) p-
value for difference =
0.006
Group A (<10 |jg/dL, n =
9) vs. Group B (>10
|jg/dL, n = 34) drawings
had the following
characteristics:
Thick and sharp lines =
33% vs. 47%; Soft lines =
78% vs. 53%%
3-337
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
With detailing = 22% vs.
29%%; Without detailing
= 78% vs. 71%%
With shading = 33% vs.
26%; Without shading =
67% vs. 82%
With distortion = 33% vs.
50%; Without distortion =
89% vs. 74%
Warm colors = 67% vs.
94%; Cool colors = 89%
vs. 65%
Rodriaues et al. (2016) Birth cohort in
Sirajdikhan and Pabna
districts
Bangladesh
2008-2011 (enrollment)
Cross-sectional
Bangladesh
n: 525 (Sirajdikhan:
Pabna: 286)
239;
Pregnant women
(gestational age <16 wk).
Blood
Child concurrent whole
blood tested using the Pb
Care II
Water samples from tube
well tested for As and Mn
during first trimester of
pregnancy and follow-up
visits at age of 1 mo, 12
mo and 20-40 mo.
Age at measurement:
20-40 mo
Cognitive development Maternal age and
using the culturally
adapted BSID-III. Age-
adjusted Z-scores
Age at outcome:
20-40 mo of age
education, child's sex,
ETS, HOME score,
maternal Raven score,
child hematocrit levels,
As, Mn.
Beta (95% CI)
Cognitive development
Pabna region3: 0.02 (SE:
0.12) per In-transformed
BLL*
Sirajdikhan regionb:
-0.02 (-0.04, 0.00)
*Unable to standardize
because P25 and median
BLL were
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Rov et al. (2011)
Chennai, India (4
representative industrial
and traffic zones)
2005-2006
Cross-sectional
School Children (3-7 yr)
n: 725
3 schools from each of
the 4 zones randomly
selected (12 schools
total); children lower and
upper kindergarten and
first grades.
Blood
Postnatal venous blood;
PbCare Analyzer
Age at measurement:
3-7 yr of age
Overall mean: 11.5
|jg/dL; mean by
genotypes: Taq A1/A1:
11.66 |jg/dL; Taq
A1/A2+A2/A2: 11.42
|jg/dL
Max: Overall: 40.5 |jg/dL
IQ using BKT (mental
age divided by
chronological age and
multiplied by 100)
Tamil-translated Binet-
Kamat Scales of
Intelligence
Age at outcome:
3-7 yr of age
Age + age 2, sex,
midarm circumference,
average monthly family
income, and family size,
parental education.
Note: stratified analysis
by genotypes (3
categories) conducted.
Beta (95% CI)b
IQ (BKT, Tamil-
translated): -4.22 (-7.10,
-1.36)
tRyqiel et al. (2021)
Mexico City
Mexico
1997-2005
Cohort
ELEMENT project
n: 85
Mother-child pairs
recruited at the Mexican
Social Security Institute
Blood
Maternal and child
venous blood; ICP-MS,
GFAAS
Age at measurement:
T1, T2, T3 (maternal);
12, 24 mo (child)
Maternal blood GM (SD):
T1: 5.27 (1.93) pg/dL
T2: 4.74 (1.96) pg/dL
T3: 4.98 (1.93) pg/dL
Infant blood GM (SD):
12 mo: 3.92 (1.80) pg/dL
24 mo: 3.49 (1.93) pg/dL
MDI assessed using
BSID-II (Spanish version)
Age at outcome: 12-24
mo
Maternal IQ (WAIS),
maternal age, infant
weight, length, SES,
infant age and sex,
current infant BLL.
Beta (95% CI) for 12-
month MDIb
T1: 0.31 (0.00, 0.62)
T2: 0.11 (-0.63, 0.86)
T3: 0.41 (-0.34, 1.17)
Beta (95% CI) for 24-
month MDIb
T1: -0.16 (-0.99, 0.66)
T2: -0.23 (-1.05, 0.59)
T3: 0.28 (-0.50, 1.06)
3-339
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Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tSanchez et al. (2011)
Mexico City
Mexico
1997-1999 (enrollment)
Followed through 24 mo
Cohort
ELEMENT study
n: 169
Mother-child pairs
recruited during
pregnancy or before
conception.
Blood
Maternal blood; ICP-MS
Age at measurement:
each trimester of
pregnancy
Mean (SD):
1st trimester (n = 139):
13.7 (3.4) |jg/dL
2nd trimester (n = 159):
24.5 (2.8) |jg/dL
3rd trimester (n = 147):
35.2 (1.9) |jg/dL
Max:
1st trimester: 20.4 |jg/dL
2nd trimester: 33.7 |jg/dL
3rd trimester: 39.0 |jg/dL.
MDI assessed using
BSID-II (Spanish version)
Scores standardized for
mother's age, mother's
IQ, duration of
breastfeeding, sex, and
weight and height z-
score at 24 mo
Age at outcome:
24 mo
Maternal age, IQ,
duration of breastfeeding,
sex, weight, and height
Z-score at 24 mo.
Beta (95% Cl)b
T1
T2
T3
-5.42 (-10.2, -0.64)
0.88 (-5.34, 7.09)
1.22 (-3.65, 6.08)
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Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Saxena et al. (2022)
Araihazar
Bangladesh
2012-2016
Cross-Sectional
Metals, Arsenic, &
Nutrition in Adolescents
study (MANAs)
n: 572
Blood
Whole blood Pb
quantified using ICP-MS.
Age at Measurement:
Mean (SD) = 14.6 (0.7)
years
Mean = 98.7 |jg/L;
Median = 91.29 |jg/L
Cognitive Effects
The Cambridge
Neuropsychological Test
Automated Battery
(CANTAB) was
administered to the
adolescents to assess
aspects of executive
function.
Linear regression models
adjusted for BMI, head
circumference, child's
years of education,
maternal intelligence
(WASI), paternal years of
education, wall type, sex,
and other blood metals -
arsenic, cadmium,
manganese, and
selenium.
Beta (95% Cl)c
Delayed Match to
sample: -3.67 (-6.59, -
0.75)
Planning: -0.05 (-0.42,
0.32)
Rapid visual processing:
-0.01 (-0.03, 0.01)
Reaction time: 0.03
(-0.01, 0.07)
Spatial recognition
memory: 1.9 (-0.88,
4.68)
Spatial span: -0.16
(-0.43, 0.11)
Spatial working memory:
1.09 (-2.54, 4.72)
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Outcome
Confounders
Effect Estimates and
95% CIs
Soetrisno and Delaado-
Saborit (2020)
West Java (Depok,
Bogor and Bekasi)
Sukatani village (control)
Indonesia
Cross-sectional
School children living in
urban locations near e-
waste facility; control site
n: 44 (22 from Bogor and
22 from Sukatani)
Children selected from
schools per teachers/
principal
recommendation.
Hair, soil, water
Hair samples from
children in Bogor and
Sukatani village. BLLs
from 36 children in Bogor
area (2010).
Age at measurement:
6-9 yr
Soil Pb mean: Depok-
Bekasi: 3653 mg/kg;
Sukatani: 93.2 mg/kg;
Water Pb: all 10 samples
below LOD; Hair Pb:
Depok-Bekasi: 0.155
mg/g; Sukatani: 0.0729
mg/kg
Max: Soil Pb: Depok-
Bekasi: 7662 mg/kg;
Sukatani: 115 mg/kg;
Hair Pb: Depok-Bekasi:
0.841 mg/g; Sukatani:
0.255 mg/kg
Academic achievement
(see also Section 3.5.1.4,
executive function)
Performance on reading,
math, writing expression
and oral language, arts,
science, social sciences,
and sports collected from
the school official alumni
report. TMT B.
Age at outcome:
6-9 yr
Age, parental education,
environmental tobacco
smoke at home, and
residential traffic
exposure.
Beta (95% Cl)d
Change in TMT-B
(seconds) per mg/g unit
of hair Pb: 54 (-3.8, 114)
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Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Sun et al. (2015)
Jiangsu Province
China
Nov 2011
Cross-sectional
Chinese National Health Blood, Urine
Research Program
n: 446
Participants recruited
from three primary
schools located in the
three towns.
Child venous blood
samples; ICP-MS
method. Morning urine
samples collected was
tested for heavy metals.
Age at measurement:
9-13 yr
GM BLL: 33.13 pg/L;
Arithmetic mean BLL:
36.99 pg/L
75th: 43.39 pg/L
90th: 56.85 pg/L
Max: 101 pg/L
IQ (CRT). Primary score Father's education,
was converted to
standard IQ scores.
Age at outcome:
9-13 yr
mother's education, BMI,
annual family income,
gender, age.
Beta (95% CI)b
-6.61 (-13.15, -0.07)
Tassiopoulos et al.
(2017)
22 PHACS clinical
research sites in the
United States, including
Puerto Rico
USA
Enrollment began in
2007; BLL data available
from 1998-2014,
developmental data
available from
1996-2010,
developmental data
available from
1996-2010
Cohort
Surveillance Monitoring
of ART Toxicities
(SMARTT)
n: 546 children with a
Bayley-111 at one year of
age who had a BPb
between 9 mo of age and
up to 3 mo after the
Bayley-111; 634 children
with a Bayley Screen at 3
yr of age and a BPb
between 9 mo of age and
up to 3 mo after the
Bayley Screen
Blood
Blood lead obtained
between the ages of 1
and 3 yr as part of
standard of care or local
guidelines are abstracted
from the medical chart
when available
Age at Measurement: 1
yr (n = 546) and 3 yr (n =
634)
Cognitive Effects
Cognition and language
neurodevelopment using
BSID-III.
Age at outcome: 1 yr
At 3 yr of age,
developmental function
was assessed with the
Bayley Screening Test
(Bayley Screen), 19 which
includes a subset of
items from the Bayley-111
with the domains of
cognition, receptive
communication,
expressive
communication, and fine
and gross motor
development.
Sex; race; ethnicity;
maternal IQ (evaluated
with the Wechsler
Abbreviated Scale of
Intelligence); maternal
education, primary
language, living
arrangement, and living
situation; household
income; geographic
region; prenatal tobacco
exposure; postnatal
tobacco exposure within
the home; age at the
developmental
evaluation; and help from
others caring for the
child.
OR (95% Cl)b (for BLLs
>=5 vs. <5 pg/dL)
Cognitive delay 1.64
(0.95, 2.90)
Receptive
communication 0.83
(0.47, 1.43)
Expressive
communication 0.91
(0.52, 1.58)
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Tuna et al. (2022)
Providence, Rhode
Island
United States
Mother-newborn
assessed for exposure
and outcome within 2 hrs
of delivery and 24 hrs of
delivery, respectively.
Cross-Sectional
Rhode Island Health
Study (RICHS)
n: 192
Placental Blood Pb
Placental Pb
concentrations quantified
using ICP-MS.
Age at Measurement:
24 hrs
Mean = 4.49 ng/g among
those with detectable Pb
BSID
Newborns' neurologic
integrity, behavioral
function, and signs of
stress assessed by NICU
Network Neurobehavioral
Scale (NNNS). Latent
Profile Analyses used to
place children in
subgroups with discrete
profiles.
Multinomial regression
models adjusted for
infant gender, maternal
age, maternal race,
maternal BMI, education
status, and smoking
status during pregnancy.
OR (95% Cl)b for
neurobehavioral profile
membership associated
with detectable Pb
(>LOD, dichotomized) vs.
Profile 2 membership
Profile 1: 0.95 (0.38,
2.35)
Profile 3: 0.97 (0.42,
2.25)
Profile 4: 0.91 (0.38,
2.20)
Profile 5: 3.42 (0.88,
13.32), p <0.1
Profile characteristics: 5
= Highest arousal,
excitability and
hypertonicity with lowest
quality of movement and
regulation (most
extreme). Other profiles:
1 = High attention and
quality of movement; 2
[referent] = Average with
lowest lethargy; 3 =
Average, required more
handling; 4 = More signs
of lethargy, hypotonicity,
nonoptimal reflexes, low
attention and arousal.
*Unstandardized due to
BLL distribution
information
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Outcome
Confounders
Effect Estimates and
95% CIs
Wan etal. (2021) n: 333
China
Cross-Sectional
Blood
Blood samples were
collected from a previous
study and analyzed for
Pb. Authors did not
report analytical method.
Age at Measurement:
Children aged 9-11 yr;
exposure group mean
(SE) = 9.93 (0.85) years;
control group (SE) = 9.62
(0.73)
FSIQ
Intelligence was tested
using the Combined
Raven's Test in China
(CRT-C2).
Multivariable linear
regression adjusted for
children's age and
gender, father's and
mother's age, education
levels and occupations,
passive smoking of the
children, and annual
family incomes.
Beta (95% Cl)a
-1.2 (-1.7, -0.60)
Median for exposure
group = 7.163 |jg/dL;
median for control group
= 3.703 |jg/dL
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Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Wang et al. (2012)
Taizhou region (Luqiao
city and Lanxi city),
Zhejiang Province (for
exposure site) and
Chun'an, Zhejiang
province (reference site)
China
June 2010
Cross-sectional
School-based study
n: 329 (Luqiao: 108,
Lanxi: 151, Chun'an: 70)
Schools located near e-
waste recycling center
and tinfoil manufacturing
area (Luqiao and Lanxi
cities). Comparison
group schools in area
dominated by agriculture
(Chun'an).
Blood, urine
Child's venous blood,
urine; ICP-MS.
Age at measurement:
11-12 yr
GM: Luqiao: 6.97 |jg/dL,
Lanxi: 8.11 |jg/dL,
Chun'an: 2.78 |jg/dL
(42%-53% had BLL >10
|jg/dL in Luqiao and
Lanxi and no one had
BLL >10 in Chun'an)
Max: Luqiao: 57.24
|jg/dL, Lanxi: 59.98
|jg/dL, Chun'an: 7.59
pg/dL
IQ (CRT) calculated from
raw score.
Age at outcome:
11-12 yr
Child's sex, birth weight,
BMI, gestation at delivery
and the mother's age at
delivery, years of
education, yearly income,
tobacco exposure during
pregnancy and alcohol
exposure during
pregnancy.
Beta (95% Cl)b
IQ (CRT)
Female: -0.097 (-0.178,
-0.016)
Male: -0.096 (-0.175,
-0.016)
AAS = atomic absorption spectrometry; ADHD = attention deficit/hyperactivity disorder; As = arsenic; BASC = Behavior Assessment System for Children; BKT = Binet Kamat Test Of
Intelligence; BLL = blood lead level; BMI = body mass index; BRIEF = Behavior Rating Inventory of Executive Functions; BSID = Bayley Scales of Infant and Toddler Development;
CANTAB = Cambridge Neuropsychological Test Automated Battery; CBLI = cumulative blood lead index; CI = confidence interval; Co = cobalt; Cr = chromium; CRS = Conners'
Rating Scales; CRT = Combined Raven's Test; Cs = cesium; Cu = copper; ELEMENT = Early Life Exposure in Mexico to Environmental Toxicants; ETS = environmental tobacco
smoke; Fe = iron; FSIQ = full-scale intelligence quotient; GFAAS = graphite furnace atomic absorption spectrometry; GM = geometric mean; Hg = mercury; HNES = Home Nurture
Environment Scale; HOME = Health Outcomes and Measures of the Environment; ICP-MS = inductively coupled plasma mass spectrometry; ICP-MS-DRC = inductively coupled
plasma mass spectrometry; K-XRF = K-shell X-ray fluorescence; LOD = limit of detection; MDAT = Malawi Developmental Assessment Tool; MDI = Mental Developmental Index; Mn
= manganese; mo = month(s); MOCEH = Mothers' and Children's Environmental Health; MSCA = McCarthy Scales of Children's Abilities; NR = not reported; OLS = ordinary least
squares; Pb = lead; PDI = Psychomotor Developmental Index; PIQ = performance intelligence quotient; PROGRESS = Programming Research in Obesity, Growth, Environment and
Social Stressors; Sb = antimony; SES = socioeconomic status; SPM = Standard Progressive Matrices; TMT = Trail Making Test; TOKS = tin, ores, kiln, smelters; VIQ = verbal
intelligence quotient; VMI = visual-motor integration WAIS = Wechsler Adult Intelligence Scale-Revised; WASI = Wechsler Abbreviated Scale of Intelligence; WISC = Weschler
Intelligence Scale for Children; WPPSI = Wechsler Preschool and Primary Scale of Intelligence; WRAT = Wide Range Achievement Test.
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bEffect estimates are not standardized because data pertaining to the BLL distribution and/or base for the log-transformation were not reported.
°Per natural log increased in centered BLLs (i.e., BLL/median).
dResults are unstandardized due to the biomarker (hair).
tStudies published since the 2013 Integrated Science Assessment for Lead.
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Table 3-7E Epidemiologic studies of Pb exposure and performance on neuropsychological tests of attention,
impulsivity, and hyperactivity, ADHD-related behaviors, and clinical ADHD in children
Referent^ and Study study Popu|atjon
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tNeuaebauer et al.
(2015)
Duisburg
Germany
2000-2002
(enrollment)
Followed through
2009-2011
Cohort
Duisburg birth
cohort study
n: 114
Pregnant women
and their offspring
Blood
Maternal venous blood; AAS
Age at measurement:
32 wk gestation (prenatal)
Mean (SD): 2.216 (1.083)
pg/dL
Med: 2.0 pg/dL
95th: 4.2 pg/dL
Max: 6.3 pg/dL
Attentional performance
using KiTAP with 5 subtests:
alertness, distractibility,
Go/No-go, divided attention,
flexibility; ADHD-associated
behavior using FBB-ADHS
Age at outcome:
Mean: 8.5 yr (KiTAP); 9.5 yr
(FBB-ADHS)
maternal
diseases, parental
lifestyle, childbirth
outcomes, HOME
Score
gMR:
KITAP
Inattention (omissions): 1.15
(1.00, 1.33)
Attention (performance
speed): 1.14 (0.98, 1.33)
FBB-ADHS
Overall ADHD: 1.061 (1.009,
1.115)
Impulsivity
1.133 (1.055, 1.216)
Hyperactivity
1.047 (0.992, 1.106)
Inattention
1.054 (0.989, 1.123)
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tEthieret al. (2015)
Arctic Quebec
(Puvirnituq)
Canada
1993-1998
(enrollment)
Followed through
2009-2020
Cohort
NCDS
n: 27
Subsample of
school-aged
children who
participated in the
Cord Blood
Monitoring Program
Blood
Cord blood and concurrent
venous blood; GFAAS
Age at measurement:
Delivery (cord); 8.6-12.6 yr old
(concurrent)
Mean (SD): 5.4 (4.1) pg/dL
Max: 17.8 pg/dL
Selective spatial attention
Visuo-spatial attention-shift
task (adapted from Posner
paradigm)
Age at outcome:
8.6-12.6 yr
Sex, age at testing
time, SES,
breastfeeding
duration, maternal
alcohol, marijuana,
cigarettes use
(Each model used a
different set of
confounders)
Beta per SD increase in In-
transformed Pb:
Cord Blood
Reaction time: 0.02b
Omission Error: -0.02b
False Alarm: 0.42 (0.08,
0.76)c
Accuracy: -0.27b
Validity Effect: -0.05b
Concurrent Blood
Reaction time: 0.52 (-0.10,
1.14)c
Omission Error: -0.10b
False Alarm: -0.16b
Accuracy: -0.17b
Validity Effect: -0.13b
tTatsuta et al. (2014) TSCD birth cohort
n: 387
Sendai, Tohoku region
Japan
Study years NR
Followed through 42
mo
Mother-infant pairs
urban areas of the
Tohoku district
Blood
Cord blood; ICP-MS.
Age at measurement:
Delivery
Median: 1.0 pg/dL
Max: 1.8 pg/dL
Sequential processing and Child sex, birth order, Betas: K-ABC
mental processing scores
(K-ABC)
Age at outcome:
42 mo
alcohol and smoking
habits, duration of
breastfeeding, annual
family income at 42
mo, and maternal IQ
(Raven SPM)
Sequential Processing:
-2.136 (-12.80, 8.531)d
Mental Processing:
(-12.41, 5.774)d
-3.319
Cohort
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Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tYorifuii et al. (2011) Faroese birth cohort Blood
Attention/working memory Child age, sex,
Faroese island
Denmark
1986-1987
(enrollment)
Followed through 7-14
yr
Cohort
n: 896 (7 yr),
(14 yr)
808
Mother-infant pairs
Cord blood; electrothermal
AAS with Zeeman background
correction.
Age at measurement:
Delivery
GM: 1.57 |jg/dL
75th: 2.2 pg/dL
assessed using WISC-R
digit span
Age at outcome:
7, 14 yr
maternal IQ (RPM),
paternal employment
and education,
maternal education,
daycare at age 7,
medical risk, and
maternal alcohol use
and smoking during
pregnancy
Beta per log-transformed
Pb:
7 yr
Digit span forward: -0.11
(-0.29, 0.07)d
<2.61 pg/g Hg: -1.70
(-3.12, -0.28)
14 yr
Digit span: -0.21 (-0.53,
0.11)d
Digit span forward: -0.04
(-0.23, 0.14)d
Digit span backward: -0.17
(-0.37, 0.04)d
<2.61 pg/g Hg: -2.73
(-4.32, -1.14)
tRuebner et al. (2019)
46 centers
United States
Study Years: NR
Followed through 1-16
yr
Cohort
CKiD Cohort study
n: 412
Children ages 1-16
yr at recruitment
with mild to
moderate CKD
Blood
Child venous blood; ICP-MS.
The BLL measurement closest
to the time of neurocognitive
testing was used for analysis
(concurrent).
Age at measurement:
NR; 2, 4, or 6 yr after study
entry
Median: 1.2 pg/dL
75th: 1.8 pg/dL
Max: 5.1 pg/dL
Attention, hyperactivity, and
response inhibition
Age-specific neurocognitive
assessments (K-CPT, CPT
III, BASC-2) administered 3,
5, 7 or 9 yr after study entry.
The last available test
results were used to
evaluate long-term effects.
Mean time between BLL
and neurocognitive testing
was 2.3 yr.
Age at outcome:
1 to >18 yr; median: 15.4 yr
Child age, sex, race,
poverty, and maternal
education
Beta: K-CPT/CPT
Attention: 1.8 (0.15, 3.45)
Adjusted BRIEF and BASC-
2 results were not reported
because they were not
statistically significant.
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Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders Effect Estil?fctfs and 95%
tRoonev et al. (2018)
Lisbon
Portugal
1997-2005
(enrollment age 8-12
yr)
Followed 7 yr age 15-
19 yr
Cohort
Casa Pia Clinical
Trial of Dental
Amalgams in
Children
n: 330
Children aged 8-12
yr at baseline in the
Casa Pia school
system
Blood
Child venous blood; flameless
AAS
Age at measurement:
8-12 yr old (baseline)
Mean (SD):
Boys: 5.26 (2.73) pg/dL
Girls: 4.42 (2.19) pg/dL
Max: 15.0 |jg/dL
Neuropsychological tests of
attention
Stroop word, Stroop color,
Stroop color/word, WISC-III
Digitspan, WAIS-III, WMS-
III, Trail Making A, Adult
Trail Making A
Age at outcome:
15-19 yr (annual
assessment for 7 yr)
Age at baseline, race,
and nonverbal IQ
(home environment,
parent's SES,
medical histories
similar across
subjects
Median beta: Boys
Stroop word: -0.118
(-0.257, 0.021)
Stroop color: -0.114
(-0.246, 0.018)
Stroop color/word: -0.117
(-0.232, -0.001)
WAIS-III digitspan: -0.049
(-0.112, 0.015)
WMS-III spatialspan: -0.012
(-0.077, 0.054)
Adult Trailmaking A: -0.02
(-0.148, 0.108)
Median beta: Girls
Stroop word: -0.01 (-0.182,
0.162)
Stroop color: 0.033 (-0.142,
0.208)
Stroop color/word: -0.019
(-0.165, 0.126)
WAIS-III digitspan: -0.06
(-0.139, 0.02)
WMS-III spatialspan: -0.019
(-0.103, 0.066)
Adult Trailmaking A: -0.165
(-0.35, 0.021)
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tChoi et al. (2020)
Seoul
Korea
Aug. 2010-Feb. 2015
Case-control
Blood
Child venous blood; GFAAS
with Zeeman background
correction
Age at measurement:
5-18 yr
Mean: 1.4 (cases) vs. 1.3
(controls) |jg/dL
ADHD status
diagnosed with K-
SADS-PL
(neuropsychological
testing by board
certified
child/adolescent
psychiatrist)
n = 355 (259
ADHD, 96 controls)
5-18 yr old patients
at a child and
adolescent
psychiatry
outpatient clinic of
Seoul National
University Hospital
Inattention and Age, sex, IQ
hyperactivity/impulsvity
assed using ADHD-RS IV
(parent rating)
Attention and executive
function assessed using
computerized SCWT and
CPT
Age at outcome:
5-18 yr
Beta direct effects: ADHD-
RS
Total ADHD severity: 2.254
(-0.278, 4.785)
Inattention: 1.053 (-0.387,
2.493)
Hyperactivity/lmpulsivity:
1.259 (-0.042, 2.560)
Beta direct effects: Conners'
CPT
Inattention (errors of
omission): 3.748 (0.091,
7.404)
Impulsivity (errors of
commission): -0.925
(-4.412, 2.562)
Response Time: 2.515
(0.013, 5.017)
Response Time Variability:
2.647 (-0.846, 6.140)
Beta direct effects: Stroop
Stroop word: -1.143
(-3.316, 1.031)
Stroop color: -0.729
(-2.832, 1.375)
Stroop color/word: 0.491
(-1.876, 2.857)
Stroop color/word
interference: 1.618 (-0.963,
4.199)
Beta interaction: Stroop
DAT1 x Pb on Inattention
(errors omission): 10.613
(-0.237, 21.463)
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Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders Effect Estil?fctfs and 95%
DAT1 x Pb on Response
Time Variability: -0.198
(-10.527, 10.132)
DRD4 x Pb on Inattention
(errors omission): -0.911
(-7.380, 5.558)
DRD4 x Pb on Response
Time Variability: -4.065
(-10.166, 2.036)
ADRA2A Mspl * Pb on
Inattention (errors omission):
2.870 (-2.340, 8.079)
ADRA2A Mspl x Pb on
Response Time Variability:
-1.588 (-6.526, 3.350)
ADRA2A Dral * Pb on
Inattention (errors omission):
5.066 (0.197, 9.934)
ADRA2A Dral x Pb on
Response Time Variability:
3.392 (-1.233, 8.017)
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Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tBoucher et al.
(2012a)
Nunavik region,
Montreal
Canada
1993-1998
(enrollment)
Followed through Sep.
2005-Apr. 2007
Cohort
Cord Blood
Monitoring Program
(CBMP)
One child from the
Environmental
Contaminants and
Child Development
Study (1996-2000)
n: 196
School children
without known
neurodevelopmenta
I disorder or
medication for
attention problems
Blood
Cord and child blood; GFAAS
with Zeeman background
correction (cord), ICP-MS
(child).
Age at measurement:
Delivery (cord), 9-13 yr (child)
Cord: 4.8 |jg/dL (mean), 3.7
|jg/dL (med); 20.9 |jg/dL (max)
Concurrent child: 2.2 |jg/dL
(mean), 2.0 |jg/dL (med), 12.8
|jg/dL (max)
Impairment in response
inhibition (Go/No-Go, ERPs
measured by EEG)
Electro-oculogram was
recorded from bipolar
miniature electrodes placed
vertically above and below
the right eye.
Age at outcome:
9-13 yr
Child age, sex, status
as adoptee; transport
by plane from remote
to larger village for
assessment; time of
assessment;
maternal age at
delivery; SES;
maternal nonverbal
reasoning abilities;
breastfeeding
duration; maternal
smoking, marijuana
use, binge drinking
during pregnancy;
docosahexaenoic
acid concentrations in
cord and child plasma
samples; Hg, PCBs
Beta per log-transformed
Pb: Cord blood
Mean Reaction Time (RT),
correct go trials: -0.05b
Mean RT, incorrect no-go
trials: -0.10b
Percent correct go trials:
-0.21 (-0.36, —0.06)c
Percent correct no-go trials:
-0.17 (-0.29, —0.05)c
Concurrent blood
Mean RT, correct go trials:
0.03b
Mean RT, incorrect no-go
trials: 0.03b
Percent correct go trials:
-0.12b
Percent correct no-go trials:
-0.16 (-0.27, —0.05)c
Rabinowitz et al.
(1992)
Taiwan
Study period NR
Cross-sectional.
N: 493
Mix of children
residing in urban or
rural environments
or near a smelter
Children grades 1-
3 recruited from
schools
Tooth
Child deciduous tooth; method
NR
Age at measurement: grades
1-3
Mean (SD): 4.6 (3.5) pg/g
Hyperactivity Syndrome
Boston Teacher
Questionnaire (BTQ)
Age at Outcome:
3
Grades 1-
Sex, # adults at
home.
Also considered
grade, child longest
hospital stay parental
education, SES, birth
outcomes,
handedness,
language at home,
and prenatal maternal
medicine, alcohol,
and smoking.
OR vs. <2.3 pg/g as
reference
2.3-7 pg/g: 1.9 (0.53, 6.5)
>7 pg/g: 2.8 (0.68, 12)
3-353
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Chandramouli et al.
(2009)
Avon
U.K
10% random
subsample of Avon
Longitudinal Study
of Parents and
Children (ALSPAC)
n = 488
Jul.-Dec. 1992 (birth) Schoo, chNdren
Followed through 8 yr
Cohort
Blood
Earlier childhood venous
blood; AAS using micro
sampling flame atomization
Age at measurement:
30 mo
Mean (SD): NR
Group 1: 0-<2 |jg/dL
Group 2: 2-<5 |jg/dL
Group 3: 5-<10 |jg/dL
Group 4: >10 pg/dL
Parent and teacher rated
hyperactivity and attention
SDQ (7 yr), Development
and Well-Being Assessment
(DAWBA) (8 yr), Test of
Everyday Attention for
Children (TEACh) (8 yr)
Age at outcome:
7-8 yr
Maternal education
and smoking, home
ownership, home
facilities score, family
adversity index,
paternal SES,
parenting attitudes at
6 mo, child sex. Also
considered child IQ.
OR for increased score:
TEACh
Group 1: reference
Group 2: 1.03 (0.66, 1.61)
Group 3: 0.99 (0.62, 1.57)
Group 4: 1.14 (0.54, 2.40)
SDQ hyperactivity
Group 1: reference
Group 2: 0.84 (0.47, 1.52)
Group 3: 1.25 (0.67, 2.33)
Group 4: 2.82 (1.08, 7.35)
tSioen et al. (2013)
Flanders
Belgium
Oct. 2002 - Dec. 2003
(enrollment)
Followed through June
2011
Cohort
Flemish Health and
Environment Study
(FLEHS 1)
n: 270
Birth cohort of
Flemish children
living in either rural
or urban areas
Blood
Cord blood, HR-ICP-MS
Age at measurement:
Delivery
median = 14.3 |jg/L
75th: 25.3 pg/L
Hyperactivity
SDQ with 5 domains:
emotional, conduct,
hyperactivity, peer and
social problems
Age at outcome:
7-8 yr
Maternal and paternal
BMI, maternal age,
weight increase of
mother during
pregnancy, smoking
during pregnancy,
smoking behavior of
maternal
grandmother before
birth of mother,
parental education,
current parental
smoking, child sex,
serious infections of
child since birth (also
tested interaction by
sex)
OR per doubling of log-
transformed Pb:
Hyperactivity: 2.940 (1.172,
7.380)d
Total difficulties: 2.167
(0.741, 6.334)d
3-354
-------
Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tFruh et al. (2019) Project Viva Blood
n: 1006
Eastern Maternal venous blood; ICP-
Massachusetts Birth cohort of MS
U.S. mother-child pairs
Age at measurement:
1999-2002 T2
(enrollment)
Followed through age Median: 1.1 pg/dL
7 yr
Parent teacher ratings of
hyperactivity using SDQ
Standardized for child age
and sex
Age at outcome:
7 yr
Maternal 2nd
trimester Hg and Mn
levels, nulliparity,
smoking during
pregnancy, IQ, and
education; paternal
education; HOME
composite score and
household income;
and child
race/ethnicity
Beta per In-transformed Pb
for hyperactivity:
SDQ-parent: 0.10 (-0.21,
0.41)
SDQ-teacher: 0.20 (-0.24,
0.64)
Cohort
tHorton et al. (2018)
Mexico City
Mexico
born 1994-2006 and
followed through age
6-16
Cohort
ELEMENT Project
n: 133
healthy, low to
moderate income
mother (18-39 yr
old)-child pairs
Teeth
tooth Pb (prenatal, postnatal
metrics derived); laser ablation
ICP-MS
Age at measurement:
tooth Pb concentration
corresponded to prenatal and
300 days after birth
Figure 1c
Externalizing behavior
(attention and hyperactivity)
BASC-2: BSI, hyperactivity
and attention symptoms
Age at outcome:
8-11 yrold
Maternal age at
delivery, maternal
education, smoking,
maternal IQ
Beta per In-transformed Pb:
Attention: 0.19 (0.02, 0.37)e
BSI (composite): 0.22 (0.06,
0.38)e
tRasnick et al. (2021) CCAAPS
n: 263
Cincinnati, OH
Born: Oct 2001—Jul
2003
Exposure: 2001-2005
Cohort
Air
LURF, air sampling at 24 sites
(C-V R2 = 0.89), predicted air
concentration at child's
residence.
Children residing >1,500 m or
<400 m from major highway
eligible.
Attention problems using
BASC-2
Age at outcome: 12 yr
Maternal education,
community-level
deprivation, blood Pb
concentrations,
greenspace, and
traffic related air
pollution.
Beta per 1 ng/m3 increase in
monthly air Pb exposure:
Attention: 0.8 (0.1, 1.5)
Median: 0.51 ng/m3 (range 0-
10.8 ng/m3)
3-355
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tLiu etal. (2014b)
Jintan, Jiangsu
province
China
Sep. 1, 2004 - Apr.
30, 2005 (age 3-5 yr)
Followed to age 6 yr
Cohort
China Jintan Child
Cohort Study
n: 1025 children
Chinese preschool
children
Blood
Venous child blood; GFAAS
Age at measurement:
3-5 yr old
Mean (SD): 6.4 (2.6) pg/dL
median = 6.0 pg/dL
75th: 7.5 pg/dL
90th: 9.4 pg/dL
Max: 32 pg/dL
Attention and ADHD
problems
CBCL (Chinese version);
Caregiver-Teacher Report
Form; normalized T scores
Age at outcome:
6 yr
Age at BLL test, sex,
preschool residence,
father's educational
level, mother's
educational level,
father's occupation,
parents' marital
status, single child
status, and child IQ
Beta:
CBCL
Attention: 0.001 (-0.002,
0.002)
ADHD: 0.136 (-0.115,
0.386)
C-TRF
Attention: 0.001 (-0.002,
0.002)
ADHD: 0.073 (-0.177,
0.322)
OR:
C-TRF
ADHD all: 1.08 (0.99, 1.18)
Boys: 1.04 (0.94, 1.16)
Girls: 1.15 (0.98, 1.35)
tWinter and Sampson PHDCN
(2017)
Chicago, Illinois
U.S.
1995-1997 (birth)
Followed through 2013
(age 17 yr)
Cohort
n: 254
Children and
caregivers living in
Chicago
Blood
Avg BLL before age 6;
methods NR
Age at measurement:
6 yr old or younger
Mean: 6.4 pg/dL
Impulsivity score
CBCL PC questionnaire
Age at outcome:
Mean: 17 yr old
Age at CBCL
assessment, sex,
race/ethnicity,
primary caregiver
immigrant
generational status,
marital status,
education level,
Temporary
Assistance for Needy
Families receipt, and
the proportion of
residential
neighborhood that is
non-Hispanic Black,
Hispanic, below the
poverty line, and
tested for Pb
exposure
Beta: CBCL
Impulsivity: 0.06 (0.005,
0.115)
3-356
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tChoi etal. (2016)
10 Cities
South Korea
2006-2010
(enrollment at 1st-2nd
grade)
Followed through age
7-9 yr
Cohort
CHEER
n: 2195
Elementary school
children
Blood
Child venous blood; AAS with
Zeeman background
correction
Age at measurement:
7-9 yr
GM: 1.56 |jg/dL
ADHD symptomology
DuPaul's ADHD rating scale
per DSM-IV
Age at outcome:
After age 7-9 yr
Age, sex, residential
area, monthly
household income,
parental marital
status, family history
of psychiatric
disorders (anxiety
disorder, ADHD,
autism and
schizophrenia),
preterm birth and
birth weight
RR (BLL>2.17 vs. <2.17
|jg/dL) for ADHD symptoms:
1.552 (1.002, 2.403)
Single parent home and BLL
>2.17 pg/dLvs. 2-parent
home and BLL <2.17 |jg/dL:
3.567 (1.595, 7.980)
3-357
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tBoucher et al.
(2012b)
Nunavik, Arctic
Quebec
Canada
1993-2000
(enrollment)
Sep. 2005-Feb. 2010
(follow-up)
Cohort
Cord Blood
Monitoring Program
and Environmental
Contaminants and
Child Development
Study
n: 279
Inuit Children
Blood
Cord and child venous blood;
AAS (cord), ICP-MS (child)
Age at measurement:
delivery (cord), 11.3 yr (child)
Mean (SD): 4.7 (3.3) pg/dL
(cord); 2.7 (2.2) pg/dL (child)
Median: 3.7 pg/dL (cord); 2.1
pg/dL (child)
Max: 20.9 pg/dL (cord); 12.8
pg/dL (child)
ADHD symptomology
assessed using the TRF
from CBCL and the DBD
rating scale
Age at outcome:
11.3 yr (average)
Child age and sex,
SES, age of the
biological mother at
birth, maternal
tobacco use during
pregnancy, and birth
weight, Hg
Cord Blood:
Attention problems Beta
(95% CI) per log-
transformed Pb: 0.05 (-0.10,
0.19)d
OR (95% CI)
ADHD inattentive type
1st fertile referent
2nd fertile 2.77 (1.00, 7.65)d
3rd fertile 2.87 (1.04, 7.94)d
ADHD hyperactive-impulsive
type
1st fertile referent
2nd fertile 0.95 (0.30, 3.00)d
3rd fertile 2.92 (1.07, 8.04)d
Child Blood:
Attention problems Beta
(95% CI) per log-
transformed Pb: 0.08 (-0.05,
0.21 )d
OR (95% CI)
ADHD inattentive type
1st fertile referent
2nd fertile 1.06 (0.42, 2.66)d
3rd fertile 1.01 (0.38, 2.64)d
ADHD hyperactive-impulsive
type
1st fertile referent
2nd fertile 4.01 (1.06,
15.23)d
3rd fertile 5.52 (1.38, 22.12)d
3-358
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tDesrochers-Couture
et al. (2019)
Nunavik, Northern
Quebec
Canada
Nov. 1993-Mar. 2002
(enrollment)
Sep. 2005-Feb. 2010
(1st follow-up)
Jan. 2013-Feb. 2016
(2nd follow-up)
Cohort
NCDS-childhood
n: 212
Inuit children from
14 coastal villages
in Nunavik,
Quebec, subsample
from the Cord Blood
Monitoring Program
and NIH-infancy
study
Blood
Cord and child venous blood;
GFAAS (cord), ICP-MS (child)
Age at measurement:
Delivery (cord), 11.4, 18.5 yr
(child)
GM (GSD): 3.80 (1.84) pg/dL
(cord); 2.34 (1.86) pg/dL
(child); 1.63 (2.00) pg/dL
(adolescent)
Median:3.73 pg/dL (cord); 2.07
pg/dL (child); 1.52 pg/dL
(adolescent)
Max: 17.80 pg/dL (cord); 12.83
pg/dL (child); 18.13 pg/dL
(adolescent)
Teacher-rated ADHD
symptomology
Teacher assessed DBD and
TRF, Achenbach's YSR,
BAARS
Age at outcome:
11.4, 18.5 yr (average)
Child age, sex, SES,
maternal age at
delivery, maternal
tobacco smoking
during pregnancy,
and birth weight
Beta (95% CI):
Child Blood:
Child externalizing behavior:
0.23 (0.08, 0.38)
Child ADHD: 0.45 (0.13,
0.78)
Direct effect: 0.09 (-0.11,
0.28)
Indirect effect: -0.02 (-0.06,
0.03)
Adolescent externalizing
behavior mediated through
child externalizing behavior
((3: 0.09, 95% CI: 0, 0.17)
tHona et al. (2015)
5 administrative
regions
South Korea
Study years NR
Case-control
n: 1001
General population
of children in 3rd to
4th grades
Blood
Child venous blood; GFAAS
with Zeeman background
correction
Age at measurement:
8-11 yr
Median: 1.81 pg/dL
75th: 2.25 pg/dL,
95th: 3.01 pg/dL
Max: 6.16 pg/dL
ADHD symptomology
Teacher/parent ratings
ADHD symptoms (ADHD-
RS); CPT
Age at outcome:
8-11 yr
Age, gender,
residential region,
paternal education
level, and yearly
income log 10-
transformed blood
Hg, Mn, urine
concentrations of
cotinine, phthalate
metabolites full-scale
IQ
Beta (95% CI) Child blood:
ADHD-RS, parent-rated
1.04 (0.18, 1.90)
ADHD-RS, teacher-rated
1.90 (0.74, 3.05)
Additionally adjusted for
FSIQ, Mn, and Hg:
ADHD-RS, parent-rated
0.68 (-0.20, 1.56)
ADHD-RS, teacher-rated
1.49 (0.32, 2.67)
3-359
-------
Referent^and Study study popu|ation Exposure Assessment
Outcome
Confounders Effect Estil?fctfs and 95%
tNiqq etal. (2016)
Michigan
United States
Study years NR
Case-control
n: 386 children, 6-
17 yr old from 267
families (148
singletons, 119
sibling pairs)
Non-ADHD: 147
ADHD: 122
Blood
Child venous blood; ICP-MS
Age at measurement: 6-17 yr
Non-ADHD: mean (SD) = 0.74
(0.35) |jg/dL
ADHD: mean (SD) = 0.94
(0.52) |jg/dL
Composite parent and
teacher ratings of ADHD
symptoms using 3 scales:
ADHD-RS: inattention and
hyperactivity-impulsivity
symptom scores
CRS-R: cognitive
(inattention) and
hyperactivity problems
subscales
Gross annual income,
HFE mutations, race,
parenting behavior,
OD/CD, Fe
hemoglobin level, sex
SWAN: inattention and
hyperactivity symptom
scores
Age at outcome: 6-17 yr
Betas of hyperactivity-
impulsivity scores perz-
score increase in Pb
modified by HFE C282Y
mutation
Parent ratings:
Mutation: 0.74 (0.52, 0.96)e
Wild-type: 0.28 (0.15, 0.41 )e
Male: 0.31 (0.14, 0.48)e
Female: 0.09 (-0.16, 0.34)e
Teacher ratings:
Mutation: 0.47 (0.22, 0.72)e
Wild-type: 0.29 (-0.04,
0.12)e
Male: 0.19 (0.07, 0.31 )e
Female: 0.11 (-0.03, 0.25)e
3-360
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tJoo etal. (2018)
Seoul, Ulsan,
Cheonan
South Korea
2006-2011
(enrollment)
Followed through 5 yr
Cohort
MOCEH
n: 575 mother-child
pairs
Pregnant women at
12-20 wk of
pregnancy in
prenatal clinics and
public health
centers
Blood
Maternal venous, cord, and
child venous blood; AAS
Age at measurement: 20 wk
(maternal); delivery (maternal
and cord); 2, 3, and 5 yr (child)
GM: Maternal 1.28 |jg/dL
(early), 1.24 (late) 0.9 (cord);
Child 1.55 (age 2), 1.43 (age
3), 1.29 (age 5)
Attention and aggressive
behavior combined
K-CBCL: Externalizing
behavior (attention and
aggressive behavior
combined);
Age at outcome:
5 yr
Maternal age at
childbirth, parity,
maternal educational
level, household
income, residential
area, and
breastfeeding
Beta (95% CI):
Externalizing behavior at 5
yr
Maternal-early pregnancy
Male: -0.72 (-3.12, 1.69)
Female: -0.45 (-2.16, 1.26)
Maternal-late pregnancy
Male: 2.99 (0.55, 5.43)
Female: 0.24 (-2.18, 2.66)
Cord blood
Male: 3.09 (-0.08, 6.26)
Female: -0.16 (-3.33, 3.01)
Child blood-2 yr
Male: 0.55 (-1.52, 2.62)
Female: 3.50 (0.97, 6.03)
Child blood-3 yr
Male: 1.13 (-1.42, 3.68)
Female: 2.05 (-1.35, 5.45)
Child blood-5 yr
(concurrent)
Male: 1.42 (-2.12, 4.95)
Female: 4.53 (-0.81, 9.86)
3-361
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tJi etal. (2018)
Boston,
Massachusetts
U.S.
1998-2013
(enrollment)
Followed through 2016
Cohort
Boston Birth Cohort
n: 299 ADHD
cases, 1180
neurotypical
controls
Mother-infant pairs
Blood
Child blood; method NR,
obtained from electronic
medical records
BLLs
Age at measurement:
<4 yr; the earlier BLL was
selected where multiple BLLs
were recorded
Mean (SD): 2.2 (1.6) pg/dL
Diagnosed ADHD
Physician-diagnosed ADHD
from electronic medical
records
Age at outcome:
Median: 6 yr
Maternal age at
delivery, maternal
race/ethnicity,
maternal education,
smoking during
pregnancy,
intrauterine infection,
parity, child's sex,
mode of delivery,
preterm birth, and
birth weight
OR (95% CI)
Continuous BLL: 1.118
(1.003, 1.247)
Categorical BLL:
2-4 vs. <2 pg/dL: 1.08
(0.81, 1.44)e
5-10 vs. <2 pg/dL: 1.73
(1.09, 2.73)e
Sex-stratified:
Girls 5-10 vs. <5 pg/dL:
0.68 (0.27, 1.69)e
Boys 5-10 vs. <5 pg/dL:
2.49 (1.46, 4.26)e
Joint Effects of sex and
BLL category:
Girls*5-10 pg/dL: 0.69
(0.28, 1.71 )e
Boys*<5 pg/dL: 3.02 (2.24,
4.06)e
Boys*5-10 pg/dL: 7.48
(4.29, 13.02)e
tParketal. (2016)
Busan
South Korea
Apr.-Sep. 2013
Case-control
n: 114 cases
(diagnosed ADHD),
114 controls
Recruitment from
child psychiatric
and pediatric clinics
from four university
hospitals
Blood
Child venous blood; GFAAS
with Zeeman background
correction
Age at measurement:
6-12 yr
GM (GSD): 1.90 (0.86) pg/dL
(cases); 1.59 (0.68) pg/dL
(controls)
Diagnosed ADHD
Diagnosed ADHD
(confirmed by [K-SADS-PL-
K]); CPT and parent-rated
ADHD symptoms among
ADHD cases
Age at outcome:
6-12 yr
Age, sex-matched OR (95% CI) per log-
controls; gestational
age, birth weight,
SES, parental
education, and
parents' smoking
behavior
transformed Pb:
ADHD total
1.60 (1.04—2.45)d
3-362
-------
RefereDCes?gnnd studV Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
n: 71 cases
(diagnosed ADHD);
58 controls
tKimetal. (2013a)
Omaha, Nebraska
U.S.
Children living near
Aug 2007-Dec 2009 a former refinery
Case-control
Blood
Child venous blood; ICP-MS
Age at measurement:
5-12 yr
GM: 1.29 |jg/dL (cases); 1.33
|jg/dL (controls); 1.65 |jg/dL
(inside Pb investigation area);
1.01 |jg/dL (outside Pb
investigation area)
ADHD
Physician-diagnosed
according to DSM-IV
Age at outcome:
5-12 yr
Matched on age, sex, OR (95% CI) per In
race and adjusted for transformed Pb
maternal smoking,
SES, and
environmental
tobacco exposure
ADHD Overall 2.52 (1.07-
5.92)e
tGeieretal. (2018) NHANES
Blood
ADD
Sex, age, SE
ES, race OR (95% CI):
n: 2109
Representative sample
U.S. Children
Child venous blood: ICP-MS
Self-reported doctor
diagnosed ADD
ADD 1.292 (1.025-1.545)
2003-2004
Age at measurement:
10-19 yr
Age at outcome:
10-19 yr
Cross-sectional
Mean (SD): 1.16 (1.27) pg/dL
3-363
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Braun et al. (2006)
Representative sample
U.S.
1999-2002
Cross-sectional
NHANES
n: 4704
Children 4-15 yr
Blood
Venous blood: GFAAS
Age at measurement:
4-15 yr old
Quintiles:
ND-0.7 |jg/dL
0.8-1.0 |jg/dL
1.1-1.3 Mg/dL
1.4-2.0 |jg/dL
679
795
857
745
Parent-reported ADHD with
prescription stimulant use
Age at outcome: 4-15 yr old
Age, sex, race,
prenatal ETS
exposure, postnatal
ETS exposure, BLLs,
preschool or
childcare attendance,
health insurance
coverage, and ferritin
levels
AOR (95% CI) Child blood:
2nd quintile (0.8-1.0): 1.1
(0.4-3.4)e
3rd quintile (1.1-1.3): 2.1
(0.7-6.8)e
4th quintile (1.4-2.0): 2.7
(0.9-8.4)e
5th quintile (>2.0): 4.1 (1.2-
14.0)e
>2.0 |jg/dL: 995
AAS = atomic absorption spectrometry; ADHD = attention deficit/hyperactivity disorder; ADHD-RS = ADHD rating scale; ADRA2A = adrenoceptor alpha 2A; AOR = adjusted odds
ratio; BAARS = Barkley Adult ADHD-IV Rating Scale; BASC = Behavior Assessment System for Children; BLL = blood lead level; BMI = body mass index; BRIEF = Behavior Rating
Inventory of Executive Functions; BSI = Behavioral Symptoms Index; CARES = Communities Actively Researching Exposure Study; CBCL = Child Behavior Check List; Cd =
cadmium; CHEER = Children's Health and Environmental Research; CI = confidence interval; CKiD = Chronic Kidney Disease in Children Study; CPT = Continuous Performance
Test; CRS-R = Conners' Rating Scale-Revised; C-TRF = Caregiver-Teacher Report Form; DAT1 = dopamine transporter; DBD = Disruptive Behavior Disorders; DRD2 = dopamine
receptor D2; DSM = Diagnostic and Statistical Manual of Mental Disorders; EEG = electroencephalogram; ELEMENT = Early Life Exposure in Mexico to Environmental Toxicants;
ERP = event-related potentials; FBB-ADHS = Fremdbeurteilungsbogen fur Aufmerksamkeitsdefizit/Hyperaktivitatstorungen; Fe = iron; FLEHS = Flemish Health and Environment
Study; GFAAS = graphite furnace atomic absorption spectrometry; GM = geometric mean; GMR = geometric mean ratio; HFE = hemochromatosis gene; Hg = mercury; HOME =
Health Outcomes and Measures of the Environment; ICP-MS = inductively coupled plasma mass spectrometry; K-ABC = Kaufman Assessment Battery For Children; K-CPT =
Conners' Kiddie Continuous Performance; KiTAP = Test of Attentional Performance for Children; K-SADS-PL-K = Kiddie Schedule for Affective Disorders and Schizophrenia Present
and Lifetime - Korean Version; LURF = Land Use Random Forest; Mn = manganese; mo = month(s); MOCEH = Mothers' and Children's Environmental Health; NCDS = Nunavik
Child Development Study; NR = not reported; OD/CD = oppositional defiant and conduct disorder; OR = odds ratio; Pb = lead; PCBs = polychlorinated biphenyls; PHDCN = Project
on Human Development in Chicago Neighborhoods; RR = relative risk; RT = reaction time; SCWT = Stroop Color-Word Test; SD = standard deviation; SDQ = Strengths and
Difficulties Questionnaire; SE = standard error; SES = socioeconomic status; SPM = Standard Progressive Matrices; SRS = Social Responsiveness Scale; SWAN = Strengths and
Weaknesses of ADHD Symptoms and Normal Behavior Scale; T1 = first trimester of pregnancy; T2 = second trimester of pregnancy; T3 = third trimester of pregnancy; TEACh =
Test of Everyday Attention for Children; TSCD = Tohoku Study of Child Development; WAIS = Wechsler Adult Intelligence Scale; wk = week(s); WMS = Weschler Memory Scale; yr
= year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 jjg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResults are unstandardized because they did not have an associated SE, CI, or p-value reported in the study.
The CI was calculated from a p-value and the true CI may be wider or narrower than calculated.
dResults are unstandardized because the log base used for exposure transformation was unspecified in the study.
eResults are unstandardized because the Pb level distribution data was not available.
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-364
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Table 3-7T
Animal toxicological studies of Pb exposure and externalizing and internalizing behaviors
Study
Species (Stock/Strain), n, Sex Timing of Exposure ^Details8 BLL QjgAdL)°rted Endpoints Examined
Externalizing Behavior
Tartaalione et al. (2020)
Rat (Wistar) GD 28 to PND 23
Oral,
PND 23:
PND 4, 7, 10, 12: Ultrasonic
Control (tap water), M/F n = 16
lactation
Vocalizations
(9/7)
In utero
0.007 |jg/mL (0.7 pg/dL)
for Control
50 mg/L, M/F, n = 16 (9/7)
0.255 pg/mL (25.5
pg/dL) for 50 mg/L
Internalizing Behavior
Corv-Slechta et al.
Mouse (C57BL/6) GD -60 to 12 mo
Oral,
PND 75 - Females:
7-12 mo: FST
(2013)
Control (distilled deionized
drinking
water) - NS, M/F, n = 8-16
water
-------
Study Species (Stock/Strain), n, Sex Timing of Exposure E^P°s.!J_re BLL f,® ^®.p°rted
Endpoints Examined
-------
Study Species (Stock/Strain), n, Sex Timing of Exposure E^P°s.!J_re BLL f,® ^®.p°rted
Endpoints Examined
for Control
9.21 ng/g (0.98 pg/dL)
for 10 |jg/mL
PND60:
0.23 ng/g (0.024 pg/dL)
for Control
0.30 ng/g (0.032 pg/dL)
for 10 pg/mL
Faulk et al. (2014)
Mouse (Agouti) GD-14toPND21
Control (distilled water), M/F, n
= 30
2.1 ppm, M/F, n = 28
16 ppm, M/F, n = 33
32 ppm, M/F, n = 29
Oral,
lactation
In utero
PND 21 (Maternal BLL):
-------
Study Species (Stock/Strain), n, Sex Timing of Exposure E^P°s.!J_re BLL f,® ^®.p°rted
Endpoints Examined
6.96 |jg/dL for 0.2%
solution
18 mo:
0.12 |jg/dL for Control
11.2 |jg/dL for 0.2%
solution
Mansouri et al. (2012)
Rat (Wistar) PND 70 to PND 100
Control (distilled water), M/F, n
= 16 (8/8)
50 mg/L, M/F, n = 16 (8/8)
Oral,
drinking
water
PND 100-Males: PND 100: OFT
2.05 |jg/dL for Control
8.8 |jg/dL for 50 mg/L
PND 100 - Females:
2.17 |jg/dL for Control
6.8 |jg/dL for 50 mg/L
Duan et al. (2017)
Mouse (CD1) PND 1 to PND 21
Oral,
PND 21: PND 7, 11, 15, 19: TST, OFT
Control (distilled water), M/F, n
lactation
= 5
16.2 |jg/L (1.6 |jg/dL) for
Control
27 ppm, M/F, n = 5
191.8 |jg/L (19.2 pg/dL)
109 ppm, M/F, n = 5
for 27 ppm
283.4 |jg/L (28.3 |jg/dL)
for 109 ppm
PND 35:
14.3 |jg/L (1.4 |jg/dL) for
Control
283.4 |jg/L (28.3 |jg/dL)
for 27 ppm
3-368
-------
Study Species (Stock/Strain), n, Sex Timing of Exposure E^P°s.!J_re BLL f,® ^®.p°rted
Endpoints Examined
376.9 |jg/L (37.7 pg/dL)
for 109 ppm
Wana et al. (2016)
Rat (Sprague Dawley)
PND 24 to PND 56 Oral,
PND 56:
PND 60-66: OFT
Control (tap water), M, n = 7
drinking
water
11 pg/L (1.1 pg/dL) for
100 ppm, M, n = 9
Control
133 pg/L (13.3 pg/dL) for
100 ppm
Shvachiv et al. (2018)
Rat (Wistar)
Intermittent Exposure: GD Oral,
PND 196:
PND 189: OFT, EPM
Control (tap water), M/F, n = 8
7 to PND 84, PND 140 to drinking
PND 196 water
<0.1 pg/dL for Control
0.2% (p/v) solution (distilled
Oral,
water), M/F, n = 9 - Intermittent
Continuous Exposure: lactation
18.8 pg/dL for 0.2%
exposure
GD 7 to PND 196 In utero
(Intermittent)
0.2% (p/v) solution, M/F, n = 9
24.4 pg/dL for 0.2%
- Continuous exposure
(Continuous)
Basha and Reddv (2015)
Rat (Wistar)
GD 6 to GD 21 In utero
PND 21:
PND 21, PND 28, 4 mo: OFT,
Control (deionized water), M, n
Hole Board Test
= 8
0.21 pg/dL for Control
0.2 % solution, M, n = 8
11.2 pg/dL for 0.2%
solution
PND 28:
0.33 pg/dL for Control
12.3 pg/dL for 0.2%
solution
4 mo:
0.19 pg/dL for Control
3-369
-------
Study Species (Stock/Strain), n, Sex Timing of Exposure E^P°s.!J_re BLL f,® ^®.p°rted
Endpoints Examined
5.9 |jg/dL for 0.2%
solution
Stansfield et al. (2015)
Rat(Long-Evans)
Control (chow), M/F, n = 11-23
1500 ppm, M/F, n = 11-23
GD Oto PND 50
Oral, diet
Oral,
lactation
In utero
PND 50:
0.6 |jg/dL for Control
22.2 |jg/dL for 1500 ppm
PND 50: Locomotor Activity
Flores-Montova and
Mouse (C57BL/6)
PND 0 to PND 28
Oral,
>PND 28 Males:
>PND 28 Hole Board Test, OFT
Sobin (2015)
Control (distilled water), M/F, n
drinking
= 19 (8/11)
water
0.2 |jg/dL for Control
Oral,
30 ppm, M/F, n = 26 (16/10)
lactation
3.93 |jg/dL for 30 ppm
230 ppm, M/F, n = 16 (12/4)
9.39 |jg/dL for 230 ppm
>PND 28 Females:
0.19 |jg/dL for Control
3.19 |jg/dL for 30 ppm
12.14 |jg/dLfor230 ppm
Neuwirth et al. (2019a)
Rat(Long-Evans)
Control (tap water), M/F, n =
NR
150 ppm, M/F, n = NR
GD Oto PND 22
Oral,
lactation
In utero
PND 22:
-------
Study
Species (Stock/Strain), n, Sex Timing of Exposure
Exposure
Details
BLL as Reported
(Mg/dL)
Endpoints Examined
-------
Study
Species (Stock/Strain), n, Sex Timing of Exposure
Exposure
Details
BLL as Reported
(Hg/dL)
Endpoints Examined
F3:
see Figure 1, n = 8-10
Sinqh et al. (2019)
Rat (Wistar) 3 mo to 6 mo
Control (distilled water), M, n =
5
2.5 mg/kg, M, n = 5
Oral, gavage
6 mo:
5.76 |jg/dL for Control
28.4 |jg/dL for 2.5 mg/kg
6 mo: EPM, Locomotor Activity
Al-Qahtani et al. (2022) Mouse (Albino) 8-9 wk to 14-15 wk Oral, gavage 14-15 wk: NR: EPM, Locomotor Activity
Control (distilled water), M, n =
10 1.2 |jg/100 mL (1.2
|jg/dL) for Control
0.2 mg/kg, M, n = 10
7.1 |jg/100 mL (7.1
|jg/dL) for 0.2 mg/kg
BLL = blood lead level; EPM = elevated plus maze; F = female; FST = forced swim test; GD = gestational day; LOD = limit of detection; M = male; MRI = magnetic resonance
imaging; mo = month(s); NR = not reported; NS = no stress; OFT = open-field test; Pb = lead; PG = pregestation; PND = postnatal day; PS = prenatal stress; TST = tail suspension
test; wk = week(s); yr = year(s).
3-372
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Table 3-8E Epidemiologic studies of Pb exposure and performance on neuropsychological tests of attention,
impulsivity, and hyperactivity, attention deficit/hyperactivity disorder-related behaviors, and
clinical attention deficit/hyperactivity disorder in children; group or population mean blood Pb
level >5 (jg/dL, any study design
Referent^and Study study Population Exposure Assessment Outcome Confounders Effect Estimates and 95% Clsa
tArbuckle et al. (2016a)
representative population
Canada
2007-2009
Cross-sectional
CHMS
n: 1080
Representative sample of
children
Blood
Child venous blood;
analytic method NR
Age at measurement:
6-11 yr old
GM: 0.90
95th: 1.96 pg/dL
ADHD symptoms
SDQ, parent-reported
ADD/ADHD
Age at outcome:
6-11 yr old
Age, sex, ORb
neonatal unit, Parent-Reported Outcomes
maternal
smoking child ADD/ADHD: 2.08 (1.01, 4.25)
age
(Supplemental Any Learning Disability: 1.41 (0.73,
2.70)
Psychotropic Medicine Taken: 2.91
(1.47, 5.79)
Table 1)
SDQ
Total Difficulties, Prenatal Smoking:
10.57 (2.81, 39.69)
Total Difficulties, No Prenatal
Smoking: 1.98 (1.41, 2.79)
Emotional Symptoms: 1.25 (0.60,
2.59)
Hyperactivity/lnattention: 2.75
(1.46, 5.16)
3-373
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Referencejmd Study study Population Exposure Assessment
Outcome
Confounders Effect Estimates and 95% Clsa
tArbuckle et al. (2016b)
representative sample
Canada
2007-2009
Cross-sectional
CHMS Blood
n: 2097
Child venous blood
Representative sample of Age at measurement:
children 6-19 yr old
ADHD symptomology
SDQ, reported ADD or
ADHD
Age at outcome:
6-19 yr old
Smoking, sex,
income
ORb
Parent or Self-Reported Outcomes,
Ages 6-19
ADD/ADHD: 2.39 (1.32, 4.32)
Learning Disability (Low Income):
0.81 (0.37, 1.81)
Learning Disability (High Income):
2.78 (1.40, 5.51)
Medicine Taken (Fasting Sample):
0.83 (0.34, 2.02)
Medicine Taken (Non-Fasting):
4.20 (1.92, 9.17)
SDQ, Ages 6-17
Total Difficulties: 2.16 (1.33, 3.51)
Emotional Symptoms: 1.08 (0.68,
1.71)
Hyperactivity/lnattention: 2.33
(1.59, 3.43)
3-374
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Referencejmd Study study Population Exposure Assessment
Outcome
Confounders Effect Estimates and 95% Clsa
tBara et al. (2018)
Montevideo
Uruguay
Cross-sectional
n: 206
Children living in areas
considered high risk for
metal exposure
Blood
Child venous blood;
flame AAS or GFAAS
Age at measurement:
5-8-year-old
4.2 |jg/dL
teacher-rated ADHD and
hyperactive behavior
CRS-R: hyperactive,
oppositional, cognitive,
and ADHD-like behaviors
(teacher ratings)
Age at outcome:
5-8-year-old
Child IQ, iron
status, and
BMI, blood Pb
testing method,
household
possessions,
maternal
education,
current parent
smoking
PRs
Cognitive Problems/Inattention
Total population (>5 vs. 5 |jg/dL):
1.02 (0.967, 1.076)
Girls: 1.01 (0.995, 1.025)
Boys: 1.01 (0.99, 1.03)
Hyperactivity
Total population (>5 vs. 5 |jg/dL):
1.01 (0.947, 1.077)
Girls: 1.02 (1, 1.04)
Boys: 0.99 (0.97, 1.01)
ADHD Index
Total population (>5 vs. 5 |jg/dL):
1.01 (0.952, 1.072)
Girls: 1.01 (0.99, 1.03)
Boys: 1.00 (0.98, 1.02)
3-375
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders Effect Estimates and 95% Clsa
tChan etal. (2015)
10 locations
U.S.
Cohort
National Institute of Child
Health and Human
Development, Study of
Early Child Care and
Youth Development
n: 266
School children
Teeth (Shed molars)
ICP-OES
Mean: 0.46 |jg/g
Disruptive behavior and
ADHD subscales
TBD completed by 3rd
grade teachers; scores
for (1) Total Disruptive
Behavior; (2) subscale
scores for ADHD,
hyperactivity/impulsivity,
inattention, and OD
Race, sex,
paternal
education,
maternal
education,
marital status,
and family SES
Change in behavior score per
|jg/g of Pb concentration in
teeth:0
DBD: -0.05
ADHD: -0.03
Impulsive: -0.06
Inattention: 0.00
Defiance: -0.09
Age at outcome:
teeth collected at 8-11 yr
old (body burden)
tFornset al. (2014)
Catalonia
Spain
Cohort
INMA
n: 385
Children of mothers
enrolled in the
population-based cohort
as part of the INMA
(Environment and
Childhood) Project
Urine
Maternal urine; ICP-MS,
values below LOD were
imputed
Age at measurement:
T1, T3
Median: 3.44, 1st; 3.63
3rd
75th: 4.64 1st, 4.84
ADHD symptoms
ADHD-DSM-IV criteria
and MSCA
Age at outcome:
4 yr old
Age, maternal Change in neuropsychological
social class, outcomes per ng/mL increase in
and maternal mother's urinary Pb
mental health concentration:
T1
GCI MSCA: 1.46 (-2.76, 5.69)
EF MSCA: 0.34 (-3.95, 4.63)
T3
GCI MSCA: -1.27 (-5.71, 3.17)
EF MSCA: -0.74 (-5.24, 3.75)
IRR:
T1
Inattention: 0.92 (0.57, 1.46)
Hyperactivity: 1.04 (0.65, 1.65)
T3
Inattention: 0.71 (0.43, 1.18)
Hyperactivity: 1.04 (0.64, 1.70)
3-376
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders Effect Estimates and 95% Clsa
tGu etal. (2018)
Wuhan
China
Case-control
Hospital based case-
control, recruitment:
n: 389 cases; 392
controls
Children 6-18 yr old
Blood
Venous sample, whole
blood, AAS
Median: 5.685 |jg/dL
ADHD, subtypes
inattention, hyperactivity
and impulsivity and
combined
Cases: ADHD DSM-IV
(subtypes defined by
inattention, hyperactivity
and impulsivity [HI] and
combined type [C]);
WISC and Parent
Symptom Questionnaire
Age at outcome:
6-18 yr old
Age, sex
(cases and
controls
compared on
IQ, maternal
alcohol,
smoking,
parental
relationship,
breastfeeding)
ORb
BLL <56.85 pg/L,
BLL >56.85 pg/L,
(1.132, 3.075)
BLL <56.85 pg/L,
(0.806, 1.954)
GG: Reference
GG: 1.865
GA/AA: 1.255
BLL >56.85 pg/L, GA/AA: 1.871
(1.014, 3.451)
tGump etal. (2017)
Upstate New York
U.S.
Cross-sectional
Environmental
Exposures and Child
Health Outcomes
n: 203
children residing in low-
to middle-income
communities
Blood
venous blood;
Age at measurement:
9-11 yr old
Externalizing behavior: Sex, race, age, Beta (95%)b
attention, impulsivity,
hyperactivity
DBD for ADHD
inattentive type and
ADHD hyperactive-
impulsive type; ASQ:I
questionnaire for ASD
(parent-rated); acute
vagal response for stress
(heart rate variability)
Age at outcome:
9-11 yr old
and SES and
Hg
ADHD-Inattention (Score) 0.01
(-0.14, 0.15)
ADHD-Hyperactivity (Score) 0.16
(0.02, 0.30)
Oppositional Defiant Disorder
(Score) 0.16 (0.02, 0.31)
3-377
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Referencejmd Study study Population Exposure Assessment
Outcome
Confounders Effect Estimates and 95% Clsa
Adjusted associations between a 1-
pg/dL increase in blood lead
Cognitive Problem/Inattention -
0.03 (-0.3, 0.2)
Hyperactivity 1.2 (0.3, 2.0)
ADHD Index 0.02 (-0.2, 0.3)
CGI Restless-Impulsive 1.2 (0.3,
2.0)
CRS-R DSM-IV
Inattentive 0 (-0.3, 0.3)
Hyperactive-Impulsive 1.1 (0.2, 2.0)
Total 0.03 (-0.2, 0.3)
tJoo et al. (2017) n: 214 cases (=19 on the
K-ARS or ADHD
Cheonan diagnosis); 214 control
South Korea (49 elementary schools)
2008-2010
Case-Control Elementary school
children
tKicinski et al. (2015)
n: 606
Blood
Sustained attention,
short-term memory,
R gender, age,
smoking,
Effect estimates between BLL and
neurobehavioral outcomes not
Flanders
Third year secondary
venous blood, ICP-MS
manual motor speed
passive
reported due lack of statistical
Belgium
school students in two
Age at measurement:
smoking,
significance
2008 and 2011
industrial areas in
13.6-17 yr old
CPT, NES
household
Cross-sectional
Flanders, Belgium
Mean: 13.8 |jg/dL
95th: 28.1 pg/dL
Age at outcome:
13.6-17 yr old
income per
capita, the
highest
occupational
category of
either parent,
and the
education level
of the mother
tHuana et al. (2016)
Mexico City
Mexico
1997-2001
Cross-sectional
ELEMENT
n: 578
Mother-child pairs
Blood
Venous blood; ICP-MS
Age at measurement:
6-13 yr old
Mean: 3.4 |jg/dL
ADHD symptomology
CRS-R, CRS-R DSM-IV
Age at outcome:
6-13 yr old
Maternal
marital status,
age,
educational
years, SES,
smoking during
pregnancy,
child's age,
sex, birth
weight.
Blood
Venous blood, AA
spectrophotometry
GM: 1.65 (cases) |jg/dL;
1.49 pg/dL (controls)
ADHD symptomology
K-ARS
Age at outcome:
6 to 10 yr old
Maternal ORb
education, AN ADHD: 1.28 (0.89, 1.83)
family history
of ADHD,
parental Inattention: 1.63 (1.03, 2.58)
marital status,
ar|d teenage Hyperactivity/impulsivity: 1.04
mother (Q ^ 2 Q7)
3-378
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Referencejmd Study study Population Exposure Assessment
Outcome
Confounders Effect Estimates and 95% Clsa
tLin etal. (2019)
n: 164
Blood, Bone
ADHD symptoms and
comorbidities
Children's age;
sex, passive
Xinhua
Children who visited a
Child venous blood; AAS
smoking (the
China
lead specialty clinic in
Tibia bone; XRF
Vanderbilt-ADHD
frequency of
Aug. 2014 - Aug. 2015
Xinhua Hospital from
Age at measurement:
Diagnostic-Parent-Rating
smoking by
Cross-sectional
August 2014-August
3-15 yr
Scale
parents and
2015
other
GM:
Age at outcome:
household
Blood:
3-15 yr
members in
the presence
Low: 4.3 |jg/dL
of children),
High: 19.6 pg/dL
parity,
Bone:
maternal
education
Low: 0.3 pg/g
levels and
High: 12.8 pg/g
family yearly
income
ORb
Inattention
BLL <10 |jg/dL: Reference
BLL >10 |jg/dL: 3.3(0.9, 12.4)
Hyperactivity/impulsivity
BLL <10 |jg/dL: Reference
BLL >10 |jg/dL: 2.0 (0.5, 7.5)
Oppositional defiant disorder
BLL <10 |jg/dL: Reference
BLL >10 |jg/dL: 2.7 (0.8, 8.9)
tLiu etal. (2014e)
n: 240
Blood
ADHD symptomology
Age and
Measures of association for blood
gender,
lead and attention outcomes not
Guiyu
Native 3-7 yr old
Child venous blood;
ADHD (H,I,C) perDSM-
residential site,
reported. Study only reports
China
kindergarten children
GFAAS
IV; CPRS-R, CTRS-R,
time, heavy
correlation analyses.
2009 -2011
who have resided in
Age at measurement:
Rutter Child Behavior
metal exposure
Cross-sectional
Guiyu for more than 2 yr
3-7 yr
Questionnaire (antisocial
after birth
behavior, neurotic)
Median: 7.33 pg/dL
75th: 9.13 pg/dL
Age at outcome:
3-7 yr
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders Effect Estimates and 95% Clsa
tLucchini et al. (2012)
Valcamonica and Garda
Lake areas in Province of
Brescia
Italy
Cross-sectional
Junior high school-age
children from 20 local
public schools
n: 299
Blood
Child venous blood;
GFAAS
Age at measurement:
11-14 yr
1.71 pg/dL, Median: 1.50
75th: 2.10 pg/dL
Max: 10.2 pg/dL
Conners'-Wells'
Adolescent Self-Report
Scale Long Form
10 subscales: family
problems, emotional
problems, conduct
problems, cognitive
problems/inattention,
anger control problems,
hyperactivity, ADHD
index, DSM-IV
(disattention), DSM-IV
(hyperactivity/impulsivity)
, and DSM-IV (Total)
Age at outcome:
11-14 yr
Sex, age at
testing,
parental
education,
SES, family
size, parity
order, BMI
Betas
Performance IQ: -1.991 (-3.918,
-0.064)
Verbal IQ: -1.863 (-3.79, 0.064)
Total IQ (Table 4): -2.237 (-4.101,
-1.372)
Total IQ (Table 5): -2.248 (-4.111,
-0.385)
tMunoz et al. (2020)
Arica
Chile
2009-2015
Cross-sectional
n: 2656
Children enrolled in a
heavy metal intervention
program
Blood
Child venous blood; AAS
Age at measurement:
3-17 yr
Median: 1.0 pg/dL
75th: 2.0 pg/dL
Parent-reported attention
deficit and hyperactivity
recorded in medical
records
Age at outcome:
3-17 yr
Age, sex,
parents' report
of children
exposure to
secondhand
tobacco
smoke,
housing
material quality
ORb
BLL >5 pg/dL: 2.33 (1.32, 4.12)
tRodriques et al. (2018) n: 225
Blood
Salvador, Bahia
Brazil
Cross-sectional
Children living near alloy GFAAS
plant Age at measurement:
7-12 yr
1.2 pg/dL
Max: 15.6 pg/dL
Child behavior
CBCL: 8 domains
including attention
Age at outcome:
7-12 yr
Sex, age,
height-for-age
Z-score,
maternal
schooling,
socioeconomic
classification,
and community
violence index,
as well as
maternal IQ
Change in Behavior (Total raw
score) per log-yg/dL decrease in
BLL:
-1.08 (-11.5, 9.3)
Change in Behavior (Total score T)
per log-ijg/dL decrease in BLL:
-0.74 (-5.3, 3.8)
3-380
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders Effect Estimates and 95% Clsa
tSkoqheim et al. (2021)
Norway
2002-2009
Case-control
Norwegian Mother,
Father and Child Cohort
Study (MoBa)
n: 397 ASD cases, 1034
controls
Children
Blood
Maternal whole blood;
ICP-SFMS atwk 17 of
gestation
Age at Measurement:
Prenatal, Week 17 og
gestation
GM (95% CI) (cases):
0.835 (0.797, 0.875)
pg/dL
GM (95% CI) (controls)
0.882 (0.860, 0.905)
pg/dL
ADHD
Diagnosis of ADHD
(NPR)
Age at outcome: 3 or less
Child sex, birth
weight, birth
year, and SGA,
maternal age
at delivery,
education,
parity, pre-
pregnancy
BMI, kg/m2),
self-reported
smoking and
alcohol intake
during
pregnancy,
FFQ-based
estimates of
seafood intake
(g/day), and
dietary iodine
intake (|jg/day)
ORb
ADHD
Q1 (Reference):
Q2
Q3
Q4
1.15 (0.87,
0.84 (0.63,
1.09 (0.82,
1.52)
1.12)
1.45)
tSobin et al. (2015)
U.S.
Cross-sectional
n: 421
Elementary school
children
Blood
Attention
2 samples 60 days apart Age at outcome:
averaged; ICP-MS or Pb 5.1-11.8 yr old
Care I
Age at measurement:
5.1—11.8 yr old
Mean: 2.7 |jg/dL (males);
2.4 |jg/dL (females)
of education
Sex, age and Beta
mother s level Motor dexterity non-dominant hand
^ 1 g3 (-1 343^ 5 2Q3)
Working memory misses: 0.11
(0.051, 0.169)
Working memory false alarms
errors: 0.05 (-0.009, 0.109)
Visual attention 5-choice movement
time (ms): 26.07 (11.331, 40.809)
3-381
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Referencejmd Study study Population Exposure Assessment
Outcome
Confounders Effect Estimates and 95% Clsa
tSoetrisno and Delaado-
Saborit (2020)
West Java (Depok,
Bogor and Bekasi)
Sukatani village (control)
Indonesia
Cross-sectional
School children living in
urban locations near e-
waste facility; control site
n: 44 (22 from Bogor and
22 from Sukatani)
Children selected from
schools per teachers/
principal
recommendation
Hair, soil, water
hair samples from
children in Bogor and
Sukatani village. BLLs
from 36 children in Bogor
area (2010).
Age at measurement:
6-9 yr
Soil Pb mean: Depok-
Bekasi: 3653 mg/kg;
Sukatani: 93.2 mg/kg;
Water Pb: all 10 samples
below LOD; Hair Pb:
Depok-Bekasi: 0.155
mg/g; Sukatani: 0.0729
mg/kg
Max: Soil Pb: Depok-
Bekasi: 7662 mg/kg;
Sukatani: 115 mg/kg;
Hair Pb: Depok-Bekasi:
0.841 mg/g; Sukatani:
0.255 mg/kg
Visual attention TMT A
Age at outcome:
6-9 yr
Age, parental
education,
environmental
tobacco smoke
at home, and
residential
traffic exposure
Change in TMT-A (seconds) per
mg/g unit of hair Pb
2.5 (-55, 60)
3-382
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders Effect Estimates and 95% Clsa
tZhana et al. (2015a)
Guangdong
China
Jan. 2012-May 2012
Cross-sectional
n: 243
Preschool children
residing near e-recycling
plant
Blood
Child venous blood,
GFAAS
Median: 7.9 |jg/dL
95th: 16.9 pg/dL
ADHD symptomology Age, sex, OR
father's work in ADHD: 2.4 (1.1, 5.2)
Parent rating per DSM-IV e-waste,
ADHD criteria Serum ferritin,
E-waste
Age at outcome: workshops
3-7 yr around the
house
ADD = attention deficit disorder; ADHD = attention deficit/hyperactivity disorder; ASD = autism spectrum disorder; ASQ:I = Ages and Stages Questionnaire Inventory; BLL = blood
lead level; BMI = body mass index; BRIEF = Behavior Rating Inventory of Executive Functions; CBCL = Child Behavior Check List; CHMS = Child Health Monitoring System; CI =
confidence interval; CPRS-R = Conners' Parent Rating Scale-reformed; CPT = Continuous Performance Test; CTRS = Conners' Teacher Rating Scale; DBD = Disruptive Behavior
Disorders; DSM = Diagnostic and Statistical Manual of Mental Disorders; ELEMENT = Early Life Exposure in Mexico to Environmental Toxicants; FFQ = Food Frequency
Questionnaire; GFAAS = graphite furnace atomic absorption spectrometry; GM = geometric mean; Hg = mercury; ICP-MS = inductively coupled plasma mass spectrometry; ICP-
OES = inductively coupled plasma optical emission spectrometry; ICP-SFMS = inductively coupled plasma sector field mass spectrometry; INMA = Infancia y Medio Ambiente
(Environment and Childhood); IQ = intelligence quotient; K-ARS = Korean ADHD Rating Scale; LOD = limit of detection; mo = month(s); NPR = Norwegian Patient Registry; NR = not
reported; OFT = open-field test; OD = oppositional defiant; Pb = lead; SDQ = Strengths and Difficulties Questionnaire; SES = socioeconomic status; SGA = small for gestational age;
TBD = to be determined; TMT A = Trail Making Test: attention; XRF = X-ray fluorescence; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bEffect estimate is unstandardized due to insufficient blood lead distribution information or insufficient information regarding log transformation.
°Result did not report confidence interval nor p-value
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-383
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Table 3-9E Epidemiologic studies of Pb exposure and externalizing behaviors including conduct disorders,
aggression, and criminal behavior in children and adolescents
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tTatsuta et al.
(2012)
Tohoku district
Japan
Study years NR
Followed through
30 mo
TSCD birth cohort
n: 306
Mother/child pairs in
Japan
Blood
Cord blood, HR-ICP-MS
Age at measurement:
delivery
Median = 1.0 |jg/dL
95th: 1.7 Mg/dL
Externalizing
behavior composite
(oppositional,
aggressive)
CBCL
Age at Outcome:
2.5 yr
Child age, birth weight, sex,
maternal age at pregnancy, delivery
type, birth order, drinking/smoking
habits in pregnancy, duration of
breastfeeding, maternal IQ,
Evaluation of Environmental
Stimulation score
Externalizing behavior
beta = -0.032b (not
significant)
Cohort
tSioen et al. (2013) Flemish Health and Blood
Flanders
Belgium
Oct. 2002-Dec.
2003 (enrollment)
Followed through
June 2011
Cohort
Environment Study
(FLEHS 1)
n: 270
Cord blood, HR-ICP-MS
Age at measurement:
delivery
Birth cohort of
Flemish children living
in either rural or urban Median = 14 3 Mg/L
areas
75th: 25.3 pg/L
Conduct problems
SDQ with 5
domains:
emotional, conduct,
hyperactivity, peer
and social
problems
Age at outcome:
7 - 8 yr
Maternal and paternal BMI, maternal
age, weight increase of mother
during pregnancy, smoking during
pregnancy, smoking behavior of
maternal grandmother before birth of
mother, parental education, current
parental smoking, child sex, serious
infections of child since birth (also
tested interaction by sex)
OR per doubling of log-
transformed Pb:
Conduct problems:
1.182 (0.319, 4.385)c
3-384
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tLiu etal. (2014b)
Jintan, Jiangsu
province
China
Sep. 1, 2004-Apr.
30, 2005 (age 3-5
yr)
Followed through
age 6 yr
Cohort
China Jintan Child
Cohort Study
n: 1025 children
Chinese preschool
children
Blood
Child venous blood; GFAAS
Age at measurement:
3-5 yr
Mean (SD): 6.4 (2.6) pg/dL
median = 6.0 pg/dL
75th: 7.5 pg/dL
90th: 9.4 pg/dL
Max: 32 pg/dL
Aggressive
behavior and
oppositional defiant
problems
CBCL (Chinese
version); Caregiver-
Teacher Report
Form; normalized T
scores
Age at outcome:
6 yr
Age at BLL test, sex, preschool
residence, father's educational level,
mother's educational level, father's
occupation, parents' marital status,
single child status, and child IQ
Parent:
Aggressive (3 (95% CI):
-0.018 (-0.264, 0.229)
Oppositional (3 (95%
CI): -0.03 (-0.28,
0.22)
Teacher:
Aggressive (3 (95% CI):
0.001 (-0.001, 0.003)
Oppositional (3 (95%
CI): 0.223 (-0.038,
0.484)
Aggressive OR (95%
CI): overall 1.07 (0.98,
1.17); boys 1.03 (0.93,
1.14); girls 1.21 (0.99,
1.47)
Oppositional OR (95%
CI): overall 1.06 (0.98,
1.15); boys 1.02 (0.92,
1.13); girls 1.12 (0.97,
1.29)
tNkomo et al.
(2017)
Soweto/Johannesb
urg
South Africa
Apr. 23-Jun. 8,
1990 (enrollment)
Followed 15-16 yr
Cohort
BT20+
n: 1322
684 females, 87.2%
Black African; 10.4%
mixed ancestry urban
residents; white and
Indian participants
excluded due to low
numbers
Blood
Child venous blood; GFAAS
with Zeeman background
correction
Age at measurement:
13 yr
Mean (SD) = 5.76 (2.42) pg/dL
median = 5.62 pg/dL
75th: 7.08 pg/dL
Max: 28 pg/dL
Violent behavior
YSR - violent
behavior
Age at outcome:
15-16 yr
Child sex, ethnicity, maternal
education, public/private hospital,
SES (unclear covariate adjustment
reporting)
physical violence (3
(95% CI): 0.05 (0.04,
0.05)
3-385
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tNkomo et al.
(2018)
Soweto/Johannesb
urg
South Africa
Apr. 23-Jun. 8,
1990 (enrollment)
Followed 14-15 yr
Cohort
BT20+
n: 1086
Black African and
mixed ancestry urban
residents; white and
Indian participants
excluded due to low
numbers
Blood
Child venous blood; GFAAS
with Zeeman background
correction
Age at measurement:
13 yr
mean (SD) = 5.6 (2.3) pg/dL
GM = 5.1 pg/dL
median = 5.4 pg/dL
Aggressive
behavior
YSR
Age at outcome:
14-15 yr
Child sex, maternal age, maternal
education at birth, marital status,
public/private hospital, SES
Direct aggression
(BLLs >10 pg/dL vs. <5
pg/dL): 0.43 (0.08,
0.78)d
tBoucher et al.
(2012b)
Nunavik, Arctic
Quebec
Canada
1992-2000
(enrollment)
2005-2010 (follow-
up)
Cohort
Cord Blood
Monitoring Program
and Environmental
Contaminants and
Child Development
Study
n: 279
Inuit Children
Blood
Cord blood; AAS
Child venous blood; ICP-MS
Age at measurement:
Avg: delivery (cord) 11.3 yr
(child)
Mean: 4.7 (cord); 2.7 (child)
Max: 20.9 (cord); 12.8 (child)
Externalizing
behavior and
OD/CD problems
CBCL andDBD
rating scale
Age at outcome:
Avg: 11.3 yr old
Child age and sex, SES, age of the
biological mother at birth, maternal
tobacco use during pregnancy, and
birth weight, Hg
Externalizing behavior
Cord: 0.09 (-0.05,
0.23 )c
Child: 0.14 (0.01,
0.26)c
OR
OD/CD
2nd vs. 1st fertile: 1.90
(0.88, 4.11 )c
3rd vs. 1st fertile: 1.53
(0.67, 3.49)c
3-386
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tBeckwith et al.
(2018)
Cincinnati, OH
U.S.
1979-1984
(enrollment)
Followed through
19-24 yr
Cohort
CLS
n: 250
Young adults from
birth cohort
Recruited pregnant
women in 1st or 2nd
trimester from inner
city neighborhoods
with historically
elevated incidence of
childhood lead
poisoning
Blood
Child venous blood; ASV
(Roda et al.. 1988))
Age at measurement:
78 mo
Mean: 7.99 |jg/dL
Max: 24.75 pg/dL
PPI score and gray
and white matter
see volume in cingulate
and ventromedial
prefrontal cortex
PPI; high resolution
anatomical MRI
with Voxel Based
Morphometry to
calculate brain
volume changes
Age at outcome:
19-24 yr old
Sex, race, age at time of imaging,
gestational age at birth, weight at
birth, maternal IQ, participant IQ,
HOME score, adult marijuana
usage, maternal prenatal alcohol
use, maternal prenatal cigarette use,
maternal narcotic use, and maternal
prenatal marijuana use
Beta
PPI
Overall: 0.22 (0.06,
0.38)e
Female: 0.16 (-0.05,
0.37)e
Male: 0.22 (-0.02,
0.47)e
tDesrochers-
Couture et al.
(2019)
Nunavik, Northern
Quebec
Canada
Nov. 1993-Mar.
2002 (enrollment)
Sep. 2005-Feb.
2010 (1st follow-up)
Jan. 2013-Feb.
2016 (2nd follow-
up)
Cohort
NCDS-childhood
n: 212
Inuit children from 14
coastal villages in
Nunavik, Quebec,
subsample from the
Cord Blood
Monitoring Program
and NIH-infancy study
Blood
Cord and venous child blood;
GFAAS (cord), ICP-MS (child)
Age at measurement:
Cord: delivery; Avg child: 11.4
and 18.5 yr
GM (GSD): 3.80 (1.84) pg/dL
(cord); 2.34 (1.86) pg/dL
(child); 1.63 (2.00) pg/dL
(adolescent)
Median:3.73 pg/dL (cord); 2.07
pg/dL (child); 1.52 pg/dL
(adolescent)
Max: 17.80 pg/dL (cord); 12.83
pg/dL (child); 18.13 pg/dL
(adolescent)
Externalizing
behavior; behavior
problems;
substance use
Behaviors -
externalizing
(CBCL),
hyperactivity-
impulsivity (DBD,
BAARS),
oppositional
defiant/conduct
disorder (DBD,
DISC); substance
use
Age at outcome:
childhood (11 yr);
adolescence (18 yr)
Child age, sex, SES, age of
biological mother at delivery,
maternal tobacco smoking during
pregnancy, birth weight, blood Hg,
house crowding, education of
primary caregiver
Beta (95% CI):
Child Blood:
Child externalizing
behavior:
0.23 (0.08, 0.38)
Direct effect on
adolescent
externalizing: 0.34
(-0.38, 1.06)
Indirect effect on
adolescent
externalizing: 0.18 (0,
0.36)
Child OD/CD: 0.37
(0.06, 0.69)
Direct effect on
adolescent CD: 0.01
(-0.10, 0.13)
Indirect effect on
adolescent CD (0.01,
-0.01, 0.03)
3-387
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tTlotlena et al.
(2022)
Johannesburg
South Africa
April-June 1990
(birth), sub-cohort
established at age
9 yr, followed
through 23-24 yr
Cohort
Young adults, n = 100 Bone
Sub-cohort (Bone
Health Cohort) of
singleton children
(born April-June
1990) from BT20
Cohort enrolling
women in 2nd and
3rd trimester residing
in Soweto-
Johannesburg
Child tibia; K-XRF
Age at measurement: NR
Mean (SD), min, med (IQR),
max: 8.7 (5.3), 0, 9 (5-12.5),
21 pg/g
Males (n = 53): 8.1 (4.4), 0, 8
(5-11), 18 pg/g
Females (n = 47): 9.4 (6.1), 0,
10 (4-14), 21 pg/g
Aggression scores
(anger, physical,
verbal, hostility)
BPAQ
Age at outcome:
23-24 yr
Age, sex, exposure to family
violence, attitude toward
neighborhood, exposure to crime
and violence in the neighborhood
Level of schooling, alcohol and drug
abuse, presence of both parents at
home, home environment, and SES
(maternal education, housing type,
participant's education/occupation)
also considered.
Beta per 1 pg/g
increase in bone Pb
Anger aggression: 0.25
(0.04, 0.37)
Physical aggression:
0.093 (-0.01, 0.27)
Verbal aggression:
0.093 (-0.05, 0.23)
Hostility: 0.03 (-0.19,
0.26)
tReuben et al.
(2019)
Dunedin
New Zealand
Apr. 1, 1972-Mar.
31, 1973
(enrollment)
Followed through
Dec. 2012
Cohort
Dunedin
Multidisciplinary
Health and
Development Study
N: 579
Birth cohort of
nationally
representative
(majority white)
children with high
rates of participation
and follow-up
Blood
Child venous blood; GFAAS
Age at measurement: 11 yr
Mean: 11.08 pg/dL
(94% above 5 pg/dL)
Antisocial behavior Sex, childhood SES, maternal IQ,
in children and family history of mental illness.
Rutter Child Scale
(averaged
parent/teacher
ratings)
Age at outcome: 11
yr
Beta
Antisocial behavior:
0.02 (0.00, 0.04)
3-388
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Chandramouli et al.
(2009)
Avon
U.K.
Jul. - Dec. 1992
(enrollment)
Followed through 8
yr
Cohort
10% random
subsample of Avon
Longitudinal Study of
Parents and Children
(ALSPAC)
n = 488
Birth cohort
Blood
Child venous blood; AAS with
micro sampling flame
atomization
Age at measurement: 30 mo
Mean (SD): NR
Group 1: 0-<2 |jg/dL
Group 2: 2-<5 |jg/dL
Group 3: 5-<10 |jg/dL
Group 4: >10 pg/dl_
Antisocial activities
Parent/teacher
ratings on
Antisocial Behavior
Interview
Age at outcome:
8 yr
Maternal education and smoking,
home ownership, home facilities
score, family adversity index,
paternal SES, parenting attitudes at
6 mo, child sex. Also considered
child IQ.
ORs for increased
score
Group 1 (0-<2 |jg/dL):
ref
Group 2 (2-<5 |jg/dL):
0.93 (0.47, 1.83)
Group 1 (5—< 10
|jg/dL): 1.44 (0.73,
2.84)
Group 1 (>10 |jg/dL):
2.90 (1.05, 8.03)
Wright et al. (2008) CLS
Cincinnati, OH
United States
1979-1984
(enrollment)
Followed through
19-24 yr
Cohort
n: 250
Young adults from
birth cohort
Recruited pregnant
women in 1st or 2nd
trimester from inner
city neighborhoods
with historically
elevated incidence of
childhood lead
poisoning
Blood
Child blood; ASV
Age at measurement: 6 yr
Median (5th-95th):
6 yr: 6.8(3.4-18) pg/dL
0-6 yr avg: 12 (6.0-26) pg/dL
Criminal arrests
County records
Age at outcome:
19-24 yr
Maternal IQ and education, sex,
Also considered potential
confounding by maternal prenatal
smoking, marijuana use, narcotic
use, and prior arrests, HOME score,
birth weight, # children in the home,
public assistance in childhood.
RRs (yes/no)
Age 6 blood Pb: 1.05
(1.01, 1.09)
Age 0-6 yr avg blood
Pb:
1.01 (0.98, 1.05)
3-389
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tFruh et al. (2019) Project Viva
n: 1006
Eastern
Massachusetts
U.S.
Birth cohort of
mother-child pairs
1999-2002
(enrollment)
Followed through
age 7 yr
Cohort
Blood
Maternal venous blood; ICP-
MS
Age at measurement:
T2
Median: 1.1 pg/dL
Parent teacher
ratings of conduct
problems using
SDQ
Standardized for
child age and sex
Age at outcome:
7 yr old
Maternal 2nd trimester Hg and Mn
levels, nulliparity, smoking during
pregnancy, IQ, and education;
paternal education; HOME
composite score and household
income; and child race/ethnicity
Parent ratings:
Overall (3 (95% CI):
0.10 (-0.10, 0.30)
Boys: 0.07 (-0.18,
0.32)
Girls: 0.13 (-0.13,
0.40)
Teacher ratings:
Overall (3 (95% CI):
0.18 (-0.08, 0.44)
Boys: 0.18 (-0.17,
0.53)
Girls: 0.17 (-0.13,
0.46)
tRuebner et al.
(2019)
46 centers
U.S.
Study Years: NR
Follow-up: NR
Cohort
CKiD Cohort study
n: 412
Children ages 1-16 yr
at recruitment with
mild to moderate CKD
Blood
Child venous blood; ICP-MS.
The BLL measurement closest
to the time of neurocognitive
testing was used for analysis
(concurrent).
Age at measurement:
NR; 2, 4, or 6 yr after study
entry
Median: 1.2 |jg/dL
75th: 1.8 |jg/dL
Max: 5.1 |jg/dL
Externalizing
behaviors,
composite index on
the BASC-2 (see
also 3.5.1 and
3.5.2)
The last available
test results were
used to evaluate
long-term effects.
Mean time between
BLL and
neurocognitive
testing was 2.3 yr.
Age at outcome:
1-16 yr
Age, sex, race, poverty, and
maternal education
Adjusted BASC-2
results were not
reported because they
were not statistically
significant.
3-390
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tNaicker et al.
(2012)
Johannesburg
South Africa
Apr. - Jun. 1990
(enrollment)
Followed through
20 yr
Birth to Twenty cohort Blood
(Bt20)
n: 1041 (487 boys,
554 girls)
Singleton children
representative of
South Africa
population
Child venous blood; GFAAS
Age at measurement:
13 yr
median = 5.4 |jg/dL; GM = 5.2
|jg/dL
Max: 28.1 pg/dL
Rule-breaking
behavior,
aggressive
behavior
YSR (adapted from
CBCL for use in
adolescents)
Age at outcome:
13 yr
SES, maternal education,
demographic factors
Attacking people -
boys (3 (95% CI): 0.54
(0.09, 0.98)d
Cohort
tRodriques et al.
(2018)
Simoes Filho,
Salvador, Bahia
Brazil
Study years NR
Cross-sectional
Simoes Filho, Brazil
n: 225
Children aged 7-12
yr, attending public in
town near ferro-Mn
alloy plant
Blood
Child venous blood; GFAAS
Age at measurement:
7-12 yr
median = 1.2 |jg/dL
Max: 15.6 |jg/dL
Behavioral
problems/disruptive
behavior
(externalizing
behavior,
aggressive
behavior, rule-
breaking behavior)
CBCL
Age at outcome:
7-12 yr
Gender, age
violence score
community
maternal IQ
Attacking people -
Adjusted total T-score
(3 (95% CI): -0.74
(-5.3, 3.8)d
3-391
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tBara et al. (2018) Montevideo sample
n: 206
Montevideo
Uruguay Children in urban
area
Study years NR
Cross-sectional
Blood
Child venous blood (fasting);
AAS with flame or graphite
furnace ionization
Age at measurement:
6-8 yr (mean = 6.75 yr)
mean = 4.2 |jg/dL
Behavior problems Child IQ, iron status, BMI, household BRIEF:
(e.g., oppositional)
CRS-R; BRIEF
Age at outcome:
6-8 yr (mean =
6.75 yr)
possessions, maternal education,
current parent smoking (also looked
at sex and Pb evaluation method in
sensitivity analyses)
Behavioral Regulation
Index (PR [95% CI]):
overall = 1.01 (1.00,
1.03); girls = 1.03
(1.00, 1.05); boys =
0.99 (0.97, 1.01)
CTRS-R:
Oppositional (PR [95%
CI]): overall = 1.00
(0.98, 1.02); girls =
1.01 (0.99, 1.04); boys
= 0.99 (0.96, 1.02)
tLiu et al. (2022b) Healthy Brains and
Behavior
Philadelphia
County, PA;
Suburbs of
Philadelphia, PA
United States
Study years NR
Cross-sectional
n: 131
Blood
Child blood; HR-ICP-MS.
Age at Measurement:
11-12 yr
Mean = 2.2 |jg/dL; Median
1.10 |jg/dL
75th: 1.8 |jg/dL
Max: 35.4 |jg/dL
Parent-report and
child-report of
externalizing
behavior
(composite)
Scores derived
from factor
analyses of 14
validated measures
of antisocial/
aggressive
behavior (RPQ,
CBCL, YSR, APSD,
CODDS, AQ from
BPAQ)
OLS regression adjusted for sex and
race.
Beta for externalizing
behavior:
Parent-reported: 0.20
(0.05, 0.34)
Child-reported: 0.20
(0.04, 0.35)
Age at outcome:
11-12 yr
3-392
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tNiqq etal. (2010) n = 326
Study location and Recruitment by
year NR
Case-control
community
advertisements,
mailings, outreach to
clinics
Blood
Child venous blood; ICP-MS
Age at Measurement: 6-17 yr
Externalizing
composite score
(oppositional and
conduct symptoms)
on parent K-SADS
Mean (SE) = 0.73 (0.04) |jg/dL Oppositional
behavior on parent
and teacher CRS
Age at Outcome:
6-17 yr
Household income, maternal
smoking, child age, sex, blood
hemoglobin, child FSIQ (WISC-IV)
Beta for SD increase in
scores per SD
increase in log-10 Pb
Parent ratings:
K-SADS externalizing
composite: 0.21 (0.05,
0.37)
CRS oppositional
behavior: 0.09 (-0.09,
0.27)
Teacher ratings:
CRS oppositional
behavior: 0.11 (-0.01,
0.23)
tAmato et al.
(2013)
Milwaukee, Wl
United States
Study years NR
Followed 7-10 yr
(blood Pb before
age 3, outcome
assessment at 4th
grade)
Wisconsin Childhood
Pb Poisoning
Prevention Project /
Milwaukee Public
School
n: 1076 unexposed;
2687 exposed; 3763
total
Exposed individuals
were more likely to be
Black or Hispanic,
and be on assisted
lunch programs
Blood
Maximum child blood Pb,
methods varied by providers
Age at measurement:
<3 yr
Mean NR; reported exposed
(BLL 10-20 |jg/dL) vs.
unexposed (<5 |jg/dL)
Max: 20 |jg/dL
School
suspensions
Unduplicated
suspension count
Age at outcome:
10 yr (4th grade)
Gender, race/ethnicity, income
(free/reduced lunch)
Suspensions
OR for exposed
(10-20 |jg/dL) vs.
unexposed (<5 |jg/dL):
2.66 (2.12, 3.32)
Cohort
3-393
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tBoutwell et al.
(2017)
106 census tracts in Blood
St. Louis City, MO
n: 59,645 children; NR
St. Louis City, MO 15,734 violent crimes Age at measurement:
United States
Study years: NR
16-year period
Other - ecological
study
St Louis residents
<72 mo age
NR
Violent crime
(crimes with
firearm, assault
crimes, robbery
crimes, homicides,
rape)
Police
department/uniform
crime report -
violent crime
(crimes with
firearm, assault
crimes, robbery
crimes, homicides,
rape)
Concentrated disadvantage; mean
age of housing; proportion occupied
by renters; domestic assaults
RR for 1% increase in
proportion of elevated
blood tests in the
census tract.
Firearm crimes: 1.03
(1.025, 1.035)
Assault: 1.03 (1.025,
1.035)
Robbery: 1.03 (1.02,
1.04)
Homicides: 1.03
(1.015, 1.045)
Rape: 1.01 (0.99, 1.03)
3-394
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tBecklev et al.
(2018)
Apr. 1, 1972-Mar.
31, 1973
(enrollment)
Followed through
38 yr
Cohort
Dunedin
Multidisciplinary
Health and
Development Study
N: 553
Birth cohort of
nationally
representative
(majority white)
children with high
rates of participation
and follow-up
Blood
Child venous blood; GFAAS
Age at measurement:
11 yr
mean = 11.01 pg/dL
Max: 31 pg/dL
Criminal offending
(criminal conviction,
recidivism,
conviction for
violent offense,
self-reported
criminal offending)
Official conviction
records from
central police
computer; self-
reported offending
interview
Age at outcome:
38 yr
Sex, age
OR (ref: no conviction)
Any criminal
conviction: 1.042 (1,
1.086)
One-time: 1.046 (0.99,
1.104)
Recidivistic: 1.039
(0.986, 1.095)
Nonviolent: 1.051
(1.003, 1.101)
Violent offense: 1.025
(0.962, 1.092)
Beta (self-report
offending)
15 yr: 0.1 (0.015,
0.185)
18 yr: 0.06 (-0.02,
0.14)
21 yr: 0.01 (-0.065,
0.085)
26 yr: 0.06 (-0.015,
0.135)
32 yr: 0.04 (-0.04,
0.12)
38 yr: 0.02 (-0.06, 0.1)
3-395
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tWriaht et al.
(2021)
Cincinnati, OH
United States
1979-1984
(enrollment)
Followed through
2013
Cohort
CLS
n: 254
Young adults from
birth cohort
Recruited pregnant
women in 1st or 2nd
trimester from inner
city neighborhoods
with historically
elevated incidence of
childhood lead
poisoning
Blood
Maternal blood (n = 219) and
child blood; ASV
Age at measurement: prenatal,
<60 mo (avg child), 60-78 mo
(avg late child), 78 mo (late
child)
Mean (SD):
Average child (0-60 mo): 14.4
(6.6) pg/dL
Total number of
arrests for each
subject (2003-2013
and lifetime),
violent crimes, drug
crimes, and
property crimes
Hamilton County
public records
Age at outcome:
18-24 yr, 27-33 yr
Birth weight (grams), maternal age
at delivery, Appearance, Pulse,
Grimace, Activity, and Respiration
scores taken at 1 min, self-reported
maternal drug use during pregnancy
that includes reports of alcohol,
marijuana and tobacco use,
maternal IQ measured by the WAIS-
R, and HOME Inventory scores
across the first 3 yr
IRR
Arrests 2003-2013 for
6-year blood Pb,
controlling for prior
arrests 1998-2003:
1.008 (0.995, 1.021)
6 yr blood Pb
Lifetime Arrests: 1.016
(1.002, 1.03)
Property Arrests: 1
(0.977, 1.023)
Drug Arrests: 1.032
(1.005, 1.06)
Violent Arrests: 1.016
(0.992, 1.039)
Adult Arrests: 1.014 (1,
1.027)
EEs for other blood Pb
sources are available
but not listed for the
sake of space.
3-396
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tEmer et al. (2020)
Milwaukee, Wl
Born June 1 1986-
Dec 31, 2005
Outcome assessed
January 1, 2005-
December 31 2015
Cohort
Adolescents enrolled
in Milwaukee public
school (2004-2016)
with BLL before age 6
N = 82,612
Capillary or venous BLLs
measured by the Milwaukee
Health department
Male:
Mean (IQR) 5.6 (5.8)
Peak (IQR): 7.0 (7.0)
Female:
Mean (IQR): 5.3 (5.0)
Peak (IQR): 6.0 (8.0)
Firearm violence
perpetration (i.e.,
coded as arrestee,
suspect or person
of interest by
Milwaukee police
department; victim
of firearm violence
assessed using
police records
Sex, race; socioeconomic status,
and year of birth
Perpetration:
OR (mean Pb): 1.03
(1.02, 1.04)
OR (peak Pb): 1.02
(1.01, 1.02)
Victimization
OR (mean Pb): 1.04
(1.03, 1.05)
OR (peak Pb): 1.02
(1.01, 1.03)
Age: < 6 yr
BASC-2 = Behavior Assessment System for Children; BLL = blood lead level; BMI = body mass index; BRIEF = Behavior Rating Inventory of Executive Functions; BT20+ = Birth to
Twenty Plus; CBCL = Child Behavior Check List; CI = confidence interval; CKD = chronic kidney disease; CKiD = Chronic Kidney Disease in Children Study; CLS = Cincinnati Lead
Study; CTRS-R = Conners' Teacher Rating Scale-Revised; DBD = Disruptive Behavior Disorder; DISC = Disrupted-in-Schizophrenia; EES = Evaluation of Environmental Stimulation;
FLEHS = Flemish Environment and Health Study; GFAAS = graphite furnace atomic absorption spectrometry; Hg = mercury; HOME = Health Outcomes and Measures of the
Environment; ICP-MS = inductively coupled plasma mass spectrometry; IQ = intelligence quotient; IQR = interquartile range; K-SADS = Kiddie Schedule for Affective Disorders and
Schizophrenia; Mn = manganese; mo = month(s); NCDS = Nunavik Child Development Study; NR = not reported; Pb = lead; PPI = Psychopathic Personality Inventory; RR = relative
risk; SDQ = Strengths and Difficulties Questionnaire; SE = standard error; SES = socioeconomic status; T2 = second trimester of pregnancy; TSCD = Tohoku Study of Child
Development; WAIS-R = Weschler Adult Intelligence Scale-Revised; yr = year(s); YSR = Youth Self-Report.
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResults are unstandardized because they did not have an associated SE, CI, or p-value reported in the study.
°Results are unstandardized because the log base used for exposure transformation was unspecified in the study.
dResults are unstandardized because the Pb level distribution data was not available.
eThe CI was calculated from a p-value and the true CI may be wider or narrower than calculated.
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-397
-------
Table 3-1OE Epidemiologic studies of Pb exposure and internalizing behaviors in children
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
Wasserman et al. (2001)
Pristina
Yugoslavia
1984-1985 (enrollment)
Followed through 1999
Cohort
N: 191
Blood
Recruitment Child blood; method NR
from prenatal
clinics
Age at outcome:
Delivery to 4-5 yr
Lifetime (to age 4-5 yr) avg
blood
Mean (SD) of log—10 Pb:
0.86 (0.12) |jg/dL, Mean:
-7.2 |jg/dL
Internalizing behavior scores
and subscores (i.e.,
anxious/depressed, somatic
complaints, and withdrawn)
assessed using maternal
ratings of CBCL
Age at Outcome: 4-5 yr
Sex, ethnicity, age, maternal
education and smoking
history, HOME score, birth
weight
Betas for log—10
change in outcome:
Internalizing
composite: 0.152
(0.023, 0.281)
Anxious/depressed:
0.041 (-0.089, 0.17)
Somatic complaints:
0.107 (-0.062,
0.276)
Withdrawn: 0.066
(-0.073, 0.205)
3-398
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
Burns et al. (1999)
Port Pirie
Australia
May 1979-May 1982
(enrollment)
Followed to age 11-13 yr
Cohort
Port Pirie
Cohort Study
(PPCS)
N: 322
Recruited
90% of live
births in a
lead smelting
community
Blood
Lifetime avg (to age 11-13
yr) blood
GM (95% CI) |jg/dL
Boys: 14.3 (13.5,
Girls: 13.9 (13.2,
15.1)
14.6)
Internalizing behavior scores
and subscores (i.e.
anxious/depressed, somatic
complaints, and withdrawn)
assessed using maternal
ratings of CBCL
Age at Outcome: 11—13 yr
Maternal age, prenatal
smoking status, IQ,
concurrent psychopathology,
and education, birth weight,
type of feeding, length of
breastfeeding, paternal
education and occupation,
birth order, family functioning,
parental smoking, marital
status, HOME score, child IQ
Beta
Male:
Internalizing
composite: 0.8 (-0.9,
2.4)b
Anxious/depressed:
0.8 (-0.2, 1.8)b
Somatic complaints:
-0.1 (-0.7, 0.4)b
Withdrawn: 0.1
(-0.4, 0.7)b
Female:
Internalizing
composite: 2.1 (0.0,
4.2)b
Anxious/depressed:
1.3 (0.1, 2.5)b
Somatic complaints:
0.3 (-0.4, 0.9)b
Withdrawn: 0.6 (0.0,
1.1)b
3-399
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
Bellinger et al. (1994b)
Boston, MA
US
1979-1980 (birth) followed
to age 8 yr
Cohort
N: 1782 Blood, Tooth
Recruitment Cord blood and shed
at birth deciduous teeth. Dentin
hospital taken from zone representing
cumulative postnatal
deposition; ASV
Age at Measurement: 6 yr
Tooth mean (SD): 3.4 (2.4)
pg/dL
Range: 0.1-28.9
10th—90th percentiles:
1.2-6.3
95th percentile: 7.4
Internalizing behavior T
scores and subscores (i.e.
anxious/depressed, somatic
complaints, and withdrawn)
assessed using teacher
ratings of CBCL
Age at Outcome: 8 yr old
Prepregnant weight, race,
Cesarean section, maternal
marital status, prenatal care,
paternal education, colic, child
current medication use,
sibship size, sex, birth weight.
Also considered potential
confounding by public
assistance, prenatal smoking,
maternal education but not
parental caregiving quality.
Betas for In-
transformed change
in internalizing T
score:
Cord blood: -0.07
(-0.23, 0.10)
Tooth: 0.43 (0.09,
0.77)
Cord blood mean (SD): 6.8
(3.1) pg/dL
Interval analyzed: 0.1-35.1
95th percentile: 12.2
tWinter and Sampson
(2017)
Chicago, Illinois
U.S.
born 1995-1997 to 2013,
followed through 17 yrold
Cohort
PHDCN
n: 254
Children and
caregivers
living in
Chicago
Blood
Child venous and capillary
blood; methods NR
Age at measurement:
before 6 yr
Avg BLL before 6 yr
Mean: 6.4 pg/dL
Internalizing on the CBCL
(i.e., anxiety and
depression); see also
Section 3.5.2 (impulsivity)
PC questionnaire)
Age at outcome:
Mean: 17 yr old
Age, sex, race/ethnicity; PC's
immigrant generational status,
marital status, education,
Temporary Assistance for
Needy Families receipt;
proportion residential
neighborhood that is non-
Hispanic Black, Hispanic, and
below the poverty line;
proportion of the child's
residential neighborhood
tested for Pb exposure
Beta
Anxiety/depression:
0.09 (0.03, 0.16)
3-400
-------
tLiu etal. (2014b)
Jintan, Jiangsu province
China
Sep. 1, 2004-Apr. 30,
2005 (age 3-5 yr)
Followed to age 6 yr
Cohort
China Jintan
Child Cohort
Study
n: 1025
children
Chinese
preschool
children
Blood
Child venous blood; GFAAS
Age at measurement:
3-5 yr old
Mean (SD): 6.4 (2.6) pg/dL
median = 6.0 pg/dL
75th: 7.5 pg/dL
90th: 9.4 pg/dL
Max: 32 pg/dL
Internalizing problems
composite and subscores
(emotionally reactive,
anxious/depressed, somatic
complaints, withdrawn, and
sleep)
CBCL (Chinese version);
Caregiver-Teacher Report
Form; normalized T scores
Age at outcome:
6 yr
Age at BLL test, sex,
preschool residence, father's
educational level, mother's
educational level, father's
occupation, parents' marital
status, single child status, and
child IQ
Internalizing
problems
Parent beta: -0.029
(-0.280, 0.222)
Teacher beta: 0.223
(-0.037, 0.484)
Teacher OR: 1.10
(1.03, 1.18)
Emotionally Reactive
Parent beta: -0.117
(-0.365, 0.131)
Teacher beta: 0.322
(0.058, 0.587)
Teacher OR: 1.10
(1.02, 1.19)
Anxiety/Depression:
Parent beta: 0.101
(-0.151, 0.354)
Teacher beta: 0.001
(-0.001, 0.003)
Teacher OR: 1.12
(1.03, 1.23)
Somatic Complaints
Parent beta: -0.171
(-0.436, 0.094)
Teacher beta: 0.001
(-0.003, 0.001)
Teacher OR: 1.01
(0.90, 1.13)
Withdrawn
Parent beta: 0.096
(-0.158, 0.349)
3-401
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
Teacher beta: 0.001
(-0.001, 0.003)
Teacher OR: 1.02
(0.93, 1.12)
tJoo etal. (2018)
Seoul, Ulsan, Cheonan
South Korea
2006-2011 (enrollment)
Followed through 5 yr
Cohort
Clinically significant
anxiety
Parent beta: 0.044
(-0.212, 0.299)
Teacher beta: 0.253
(0.016, 0.500)
Teacher OR: 1.12
(1.03, 1.23)
MOCEH
n: 575
mother-child
pairs
pregnant
women
Blood
Maternal venous blood, cord
blood, and child blood; AAS
Age at measurement: 20 wk
gestation (maternal); delivery
(cord); 2,3 and 5 yr (child)
GM:
Maternal 1.28 |jg/dL (early),
1.24 (late) 0.9 (cord)
Child 1.55 (age 2), 1.43 (age
3), 1.29 (age 5)
Internalizing behavior
K-CBCL (emotional
reactivity, anxious/
depressed, somatic
complaints, and
withdrawn/depressed states)
See also Section 3.5.2
Age at outcome:
5 yr old
Maternal age at childbirth,
parity, maternal educational
level, household income,
residential area, and
breastfeeding
Beta (95% CI):
Internalizing at 5 yr
Maternal-early
pregnancy
Male: -0.16 (-2.54,
2.23)
Female: -0.13
(-1.86, 1.60)
Maternal-late
pregnancy
Male: 2.55 (0.22,
4.88)
Female: -0.18
(-2.66, 2.31)
Cord blood
Male: 2.44 (-0.74,
5.63)
Female: -1.00
(-4.30, 2.29)
3-402
-------
Re'erenDCifgn„d S,Udy p4SS8o„ Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
Child blood-2 yr
Male: -0.03 (-2.07,
2.00)
Female: 2.94 (0.36,
5.52)
Child blood-3 yr
Male: 0.25 (-2.33,
2.82)
Female: 2.76 (-0.73,
6.26)
Child blood-5 yr
(concurrent)
Male: 1.23 (-2.10,
4.56)
Female: 5.65 (0.50,
10.80)
tFruh etal. (2019)
Eastern Massachusetts
U.S.
1999-2002 (enrollment)
Followed through age 7 yr
Cohort
Project Viva Blood
n: 1006
Maternal venous blood; ICP-
Birth cohort of MS
mother-child Age at measurement:
pairs T2
Median: 1.1 pg/dL
Parent teacher ratings of
emotional problems
SDQ
Standardized for child age
and sex
Age at outcome:
7 yr old
Maternal 2nd trimester Hg
and Mn levels, nulliparity,
smoking during pregnancy,
IQ, and education; paternal
education; HOME composite
score and household income;
and child race/ethnicity
Parent ratings:
Overall (3 (95% CI):
0.30 (0.05, 0.55)
Boys: 0.17 (-0.17,
0.50)
Girls: 0.52 (0.18,
0.86)
Teacher ratings:
Overall (3 (95% CI):
0.07 (-0.22, 0.35)
Boys: 0.02 (-0.33,
0.37)
Girls: 0.12 (-0.31,
0.54)
3-403
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
tSioen et al. (2013)
Flanders
Belgium
Oct. 2002 - Dec. 2003
(enrollment)
Followed through June 2011
Cohort
Flemish
Health and
Environment
Study
(FLEHS 1)
n: 270
Birth cohort of
Flemish
children living
in either rural
or urban
areas
Blood
Cord blood, HR-ICP-MS
Age at measurement:
delivery
median = 14.3 |jg/L
75th: 25.3 pg/L
Emotional problems
SDQ with 5 domains:
emotional, conduct,
hyperactivity, peer and social
problems
Age at outcome:
7 - 8 yr
Maternal BMI, age at
pregnancy, weight increase
during pregnancy, smoking,
paternal BMI, if parents
smoke, smoking behavior
maternal grandmother before
the birth of the mother,
parental education, child sex,
serious child infections
OR per doubling of
log-transformed Pb:
Emotional problems:
0.900 (0.524, 1.547)b
tRokoff et al. (2022)
New Bedford, MA
Born: 1993-1998
Cohort
Children
residing near
Superfund
site
n: 468 of 788
mother-infant
pairs.
Blood
Cord blood; ICP-MS
Child blood; medical records,
method NR
Age at measurement:
delivery
Mean (SD) cord BLL
CPRS: 1.37 pg/dL (0.94)
BASC-2: 1.37 pg/dL (0.95)
Mean (SD) peak postnatal
BLL
CPRS: 6.68 pg/dL (3.95)
BASC-2: 6.58 pg/dL(3.87)
Internalizing Behavior
Anxiety, Depression,
Somatization, and
Internalizing Problems on
BASC-2 SRP
Anxious-Shy and
Psychosomatic on CPRS
Anxious-Shy on CTRS
Age(s) at outcome: 8 yr
(CPRS and CTRS) and 15-
years (BASC-2)
Maternal age, marital status,
parity, parental education),
household income, maternal
smoking, alcohol consumption
during pregnancy, pre-
pregnancy weight, height, and
gestational weight gain, BMI,
prenatal social disadvantage
index, HOME score, maternal
IQ
No interactions between
chemicals; linear regression
models adjusted for Mn and
organochlorines,
BASC-2 SRP
Anxiety: 1.78 (0.58,
2.99)
Depression: 0.79
(-0.39, 1.97)
CPRS
Psychosomatic
Boys: 2.08 (0.07,
4.10)
Girls: 0.48 (-1.00,
1.97)
C-R functions
presented
3-404
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
tRasnick et al. (2021)
CCAAPS
Air
Internalizing and
Maternal education,
B (anxiety score) =
n: 263
Externalizing Behavior
community-level deprivation,
3.1 (95% CI: 0.4,
Cincinnati, OH
LURF, air sampling at 24
blood Pb concentrations,
5.7) per ng/m3
sites (C-V R2 = 0.89),
BASC-2; internalizing
greenspace, and traffic
Note: no association
Born: Oct2001-Jul 2003
predicted air concentration at
behaviors (anxiety,
related air pollution.
with depression,
Exposure: 2001-2005
child's residence.
depression, somatization),
externalizing behaviors
somatization,
conduct problems,
Cohort
Children residing >1,500 m or
(aggression, conduct
hyperactivity,
<400 m from major highway
problems, and hyperactivity),
withdrawal behaviors
eligible.
behavioral symptoms index
(attention problems,
Median: 0.51 ng/m3 (range 0-
atypicality, and withdrawal)
10.8 ng/m3)
Age at outcome: 12 yr
tRuebner et al. (2019)
CKiD Cohort
Blood
Internalizing behaviors,
Age, sex, race, poverty, and
Adjusted BASC-2
study
composite index on the
maternal education
results were not
46 centers
n: 412
Child venous blood; ICP-MS.
BASC-2 (see also 3.5.1 and
reported because
U.S.
Children with
The BLL measurement
closest to the time of
3.5.2)
they were not
statistically
Study Years: NR
mild to
neurocognitive testing was
The last available test results
significant.
Follow-up: 1-16 yr
moderate
used for analysis
were used to evaluate long-
CKD
(concurrent).
term effects. Mean time
Cohort
between BLL and
Age at measurement:
NR; 2, 4, or 6 yr after study
entry
Median: 1.2 |jg/dL
75th: 1.8 |jg/dL
Max: 5.1 |jg/dL
neurocognitive testing was
2.3 yr.
Age at outcome:
1-16 yr
3-405
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
tHorton et al. (2018)
ELEMENT
Tooth
Internalizing behavior on the
Maternal age at delivery,
Project
BASC-2. See Section 3.5.2
maternal education, smoking
Mexico City
n: 133
Tooth Pb (prenatal, postnatal
(attention and hyperactivity)
SES, maternal IQ
Mexico
metrics derived); laser
and BSI
born 1994-2006 and
healthy, low
ablation ICP-MS
followed through age 6-16
to moderate
Age at measurement:
Age at outcome:
Cohort
income
tooth Pb concentration
8-11 yrold
mother (18-
corresponded to prenatal and
39 yr old)-
300 days after birth
child pairs
Figure 1c
Beta: BASC-2
Internalizing (10 mo)
NR
Anxiety (12 mo)
(95% CI NR)C
0.4
Internalizing
composite result was
not reported because
it was not statistically
significant.
tDohertv et al. (2020)
New Hampshire
U.S.
2009 to 2014-2019
Cohort
NHBCS
n: 371
(SRS-2); 318
(BASC-2)
Mother-child
pairs
Toenails
Maternal and infant toenails;
Median (maternal prenatal):
0.14 |jg/g (SRS-2), 0.13 pg/g
(BASC-2); Median (maternal
postnatal): 0.10 pg/g (SRS),
0.11 pg/g (BASC-2); Median
(infant): 0.35 pg/g (SRS-2),
0.37 pg/g
Internalizing Behaviors on
the BASC-2; see also
Section 3.5.2.2
Age at outcome:
3 yr
Maternal age, maternal BMI,
parental education, maternal
smoking, marital status,
parity, child age at last
breastfeeding, Healthy Eating
Index score, year of birth, sex,
and age of the child at testing
Exposure was log2
transformed.
Betas per 1 pg/g
increase in toenail
Pb concentration.
Total
Maternal prenatal:
-0.14 (-0.28, 0.00)d
Maternal postnatal:
0.06 (-0.05, 0.18)d
Child: 0.01 (-0.14,
0.16)d
Males
Maternal prenatal:
-0.17 (-0.36, 0.01 )d
Maternal postnatal:
0.31 (0.15, 0.47)d
Child: 0.09 (-0.13,
0.31 )d
Females
3-406
-------
Re,ereDCi?gn„d S,Udy Population Exposure Atsessmen. Outeome
Maternal prenatal:
-0.16 (-0.33, 0.01 )d
Maternal postnatal:
-0.04 (-0.20, 0.13)d
Child: -0.15 (-0.36,
0.06)d
BASC = Behavioral Assessment System for Children; BLL = blood lead level; BMI = Body Mass Index; BRIEF = Behavior Rating Inventory of Executive Functions; CBCL = Child
Behavior Check List; CCAAPS = Cincinnati Childhood Allergy and Air Pollution Study; CI = confidence interval; CKD = chronic kidney disease; CKiD = Chronic Kidney Disease in
Children Study; CPRS = Conners' Parent Rating Scale; C-TRF = Caregiver-Teacher Report Form; CTRS = Conners' Teacher Rating Scale; C-V R2 = cross validated R-square; DSM
= Diagnostic and Statistical Manual of Mental Disorders; ELEMENT = Early Life Exposure in Mexico to Environmental Toxicants; FSIQ = full-scale intelligence quotient; GFAAS =
graphite furnace atomic absorption spectrometry; ICP-MS = inductively coupled plasma mass spectrometry; HOME = Health Outcomes and Measures of the Environment; IQ =
intelligence quotient; K-CBCL = Korean Child Behavior Check List; LURF = Land Use Random Forest; MOCEH = Mothers' and Children's Environmental Health; NHBCS = New
Hampshire Birth Cohort Study; NR = not reported; OR = odds ratio; Pb =lead; PC = primary caregiver; PHDCN = Project on Human Development in Chicago Neighborhoods; SDQ =
Strengths and Difficulties Questionnaire; SRP = Self-Report of Personality; SRS = Social Responsiveness Scale; yr = year(s); T2 = second trimester of pregnancy.
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResults are unstandardized because the log base used for exposure transformation was unspecified in the study.
°Results are unstandardized because they did not have an associated SE, CI, or p-value reported in the study.
dResults are unstandardized because the biomarker used for Pb exposure measurement is toenails.
tStudies published since the 2013 Integrated Science Assessment for Lead.
~ f . Effect Estimates
Confounders gnd 95% C|j.a
3-407
-------
Table 3-11E Epidemiologic studies of Pb exposure and motor function in children
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Ris et al.
(2004)
Cincinnati, OH
US
1979-1985
(enrollment)
Cohort
Cincinnati Lead Blood
Study (CLS)
N: 195
Visuoconstructio Maternal IQ, SES,
Birth cohort
recruited
prenatally from
obstetrical
clinics
Prenatal maternal blood: NR
Average Childhood blood (mean of 20 quarterly
concentrations obtained from 3-60 mo): NR
78 mo blood lead: NR
n (Block Design
Subtest, ROCF-
Accuracy) and
Fine-Motor
(Grooved
Pegboard Test,
Finger Tapping
Test) factors
Age at outcome:
15-17 yr
total average HOME
scores,
and adolescent
marijuana
consumption
Beta
Visuoconst ruction
Prenatal: -0.157 (-0.277,
-0.037)b
Average: 0.028 (-0.052, 0.108)b
78-month: 0.014 (-0.088, 0.116)b
Fine-motor
Prenatal: -0.017 (-0.056, 0.022)b
Average: -0.016 (-0.041, 0.009)b
78-month: -0.046 (-0.077,
—0.015)b
Bhattacharva
et al. (1995)
Cincinnati, OH
US
1979-1984
(enrollment)
Cohort
Cincinnati Lead Blood
Program Project
N: 202
Pregnant
mothers living in
older houses in
poor condition
and with
chipping lead-
based paint and
lead laden dust
GM (SD) (min-max) ug/dL:
Prenatal maternal blood: 8.0 (1.58) (2-22)
Average Childhood blood (mean of 20 quarterly
concentrations obtained from birth to 5 yr): 11.9
(1.5) (4-28)
Postural
balance,
including sway
area (SA) and
sway length (SL)
Age at outcome:
5 yr
Age, height, weight,
birth length, birth
weight, Hgb, TIBC,
Minimum middle-ear
pressure, smoking
during pregnancy,
HOME score at 36
mo, foot area, sports
participation, race,
known occurrences of
bilateral otitis media
Betas
Eyes open
SA: 0.059 (0.024, 0.093)
SL: 0.145 (0.088, 0.201)
Eyes closed
SA: 0.043 (0.01, 0.076)
SL: 0.121 (0.069, 0.173)
Eyes open, foam
SA: 0.046 (-0.175, 0.266)
SL: 0.113 (0.065, 0.16)
Eyes closed, foam
SA: 0.055 (0.018, 0.091)
SL: 0.86 (0.277, 1.443)
3-408
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Dietrich et al. N: 245
(1993)
Cincinnati, OH
US
1979-1984
(enrollment)
Cohort
Pregnant
mothers living in
older houses in
poor condition
and with
chipping lead-
based paint and
lead laden dust
Blood
Maternal and child venous blood
Age at measurement: T1 (maternal), delivery, 1
2, 3, 4, 5, 6 yr (child)
Mean (SD) (min-max) ug/dL:
T1: 8.4 (3.8) (1-27)
Neonatal: 4.8 (3.1) (1-22)
1 yr
2 yr
3 yr
4 yr
5 yr
6 yr
10.5 (4.9) (3-35)
17.1 (8.3) (6-49)
16.2 (7.6) (4-50)
14.0 (7.1) (4-45)
11.9 (6.4) (3-38)
10.1 (5.6) (2-33)
Bilateral
coordination,
visual-motor
contol, upper-
limb speed and
dexterity, and
fine motor
composite
assessed using
BOTMP
Age at outcome:
6 yr
NR
Beta
Bilateral coordination
Prenatal: -0.04 (-0.197, 0.117)b
Neonatal: -0.15 (-0.326, 0.026)b
Average (3-60 mo): -0.11
(-0.188, -0.032)b
Concurrent: -0.18 (-0.258,
—0.102)b
Visual-motor control
Prenatal: 0.06 (-0.097, 0.217)b
Neonatal: -0.1 (-0.296, 0.096)b
Average (3-60 mo): -0.05
(-0.148, 0.048)b
Concurrent: -0.12 (-0.218,
-0.022)b
Upper-limb speed and dexterity
Prenatal: -0.2 (-0.435, 0.035)b
Neonatal: -0.45 (-0.724,
—0.176)b
Average (3-60 mo): -0.19
(-0.327, -0.053)b
Concurrent: -0.31 (-0.447,
—0.173)b
Fine motor composite
Prenatal: -0.14 (-0.552, 0.272)b
Neonatal: -0.49 (-0.96, -0.02)b
Average (3-60 mo): -0.28
(-0.515, -0.045)b
Concurrent: -0.46 (-0.715,
-0.205)b
3-409
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Wasserman et
al. (2000)
K. Mitrovica
and Pristina
Kosovo,
Yugoslavia
1985-1986
(enrollment)
Followed 54
mo
Cohort
Yugoslavia
Prospective
Study
N: 283 children
Pregnant
women
recruited from
K. Mitrovica
(lead smelter,
refinery, and
battery factory)
and Pristina
(town 40 km
south)
Blood
Prenatal maternal, delivery, and subsequent 6-
month interval venous blood: NR
Fine motor and
gross motor
composites
assessed using
BOTMP
Visual motor
integration
assessed using
Beery Test of
VMI
Age at outcome:
54 mo
Maternal age,
parental education,
number of siblings,
living arrangement,
HOME score at 3 yr,
maternal intelligence
at 2 yr (RSPM),
birthweight, BMI at 54
mo, child sex,
opportunities to
practice motor skill,
incomplete
lateralization
Beta for log—10 transformed Pb
Fine motor composite: -0.17
(-1.503, 1.163)c
Gross motor composite: 0.03
(-1.538, 1.598)°
VMI: -0.24 (-0.632, 0.152)c
tKim et al.
MOCEH study
(2013c) and
n: 884
Kim et al.
(2013b)
Mothers
recruited before '
Seoul,
20th wk of
Cheonan and
pregnancy
Ulsan
between and
Korea
were in
2006-2010
locations
Followed 6 mo
(Seoul,
Cohort
Cheonan and
Ulsan)
Blood
Maternal blood samples measured for Pb, Cd in
early (<20 wk) pregnancy and late (med = 39
wk) pregnancy
Age at measurement:
Early and late pregnancy
Early pregnancy: 1.4 (GM), 2.1 (90th), 9.8 (max)
pg/dL
Late pregnancy: 1.3 (GM); 2.1 (90th), 4.3 (max)
pg/dL
GM also available separately by 3 sites
PDI assessed
using BSID-II
(Korean version)
Age at outcome:
6 mo
Birth weight, infant
sex, maternal age
and education, family
income, breastfeeding
status, residential
area.
Beta
Early: 0.28 (
Late: -1.38 i
¦1.19, 1.75)
-3.31, 0.55)
Early:
Cd <1.47 |jg/L: 2.70 (0, 5.39)
Cd >1.47 |jg/L: -1.17 (-3.27,
0.94)
Late:
Cd <1.51 |jg/L: 0.18 (-2.70, 3.07)
Cd >1.51 |jg/L: -2.86 (-5.55,
-0.16)
3-410
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tKim et al.
(2018b)
4 cities: Seoul,
Anyang, Ansan
and Jeju
Korea
2011-2012
(enrollment)
Followed
through 24 mo
Cohort
CHECK cohort
n: 140
birth cohort-
pregnant
women
recruited from 4
cities in Korea
before delivery
Blood
Maternal blood; method NR
Age at measurement: delivery
Median (IQR):
Maternal: 2.7 (3.5, 5.7) pg/dL
Cord: 1.2 (0.8, 1.7) pg/dL
PDI assessed
using BSID-II
(Korean version)
Age at outcome:
13-24 mo
BPA, and phthalates,
maternal age
(continuous), birth
delivery mode
(categorical), monthly
household income
(categorical), child's
sex, and BDI
(continuous) of the
mother, gestational
age (continuous),
primiparous
(categorical), and
pre-pregnancy BMI
(categorical)
Beta (maternal blood)
Overall: -15.45 (-30.12, -0.79)
Boys: -18.32 (-45.35, 8.71)
Girls:-7.48 (-42.10, 27.15)
tY Ortiz et al.
(2017)
Mexico City
Mexico
Jul 2007-Feb
2011
Followed
through 24 mo
PROGRESS
birth cohort
n: 536
Women <20 wk
of gestation and
planning to
reside in Mexico
City for the next
3 yr.
Blood
Maternal blood analyzed using ICP-MS.
Age at measurement:
T2, T3
Mean:
T2: 3.7 pg/dL
T3: 3.9 pg/dL
Motor
development
assessed using
BSID-II I.
Standardized
scores (mean:
100, SD: 15).
Age at outcome:
24 mo
Infant sex, birth
weight, gestational
age, maternal age,
maternal IQ (WAIS
Spanish version),
HOME score.
Beta for log-transformed Pb
Motor Development:
T2: 1.97 (-2.46, 6.40)b c
T3: -11.01 (-17.55, -4.48)bc
Cohort
3-411
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tLiu et al.
Birth cohort
(2014c)
from 3 medical
centers
Pearl River
n: 362 mother-
Delta Region,
infant pairs (141
Guangdong
high Pb group
China
with cord BLL
>3.92 pg/dL and
Jan 2009-Jan
102 low Pb
2010
group <1.89
(enrollment)
pg/dL)
Followed for 3
yr
Cohort
tRvaiel et al.
ELEMENT
Blood
Cord blood and child blood analyzed using
GFAAS. 2 exposure groups created based on
cord BLL below 25th percentile (low) and above
the 75th percentile (high). Age at measurement:
At birth (cord), 6, 12, 24 and 36 mo (postnatal
child)
High and low Pb groups: cord BLLs: 5.63 and
1.35 |jg/dL; 6 mo BLL: 4.03 and 2.85 pg/dL; 12
mo: 4.87 and 3.79 pg/dL; 24 mo: 4.39 and 3.31
pg/dL; 36 mo: 3.94 and 3.28 pg/dL
PDI assessed
using BSID-II
(Chinese
version)
Age at outcome:
36 mo
Birth weight, sex,
maternal education,
IQ (WISC-R),
hemoglobin level,
smoking, age,
parental occupations,
household annual
income, HNES total
score
Beta comparing PDI score at 36
mo in high exposed (Cord BLL
>3.92 pg/dL) vs. low exposed
(Cord BLL <1.89 pg/dL):
-1.302 (-1.572, -1.031)
(2021)
Mexico City
Mexico
1997-2005
Cohort
project
n: 85
Mother-child
pairs recruited
at the Mexican
Social Security
Institute
Blood
Maternal and child venous blood; ICP-MS,
GFAAS
Age at measurement:
T1, T2, T3 (maternal); 12, 24 mo (child)
Maternal blood GM (SD):
T1
T2
T3
PDI assessed
using BSID-II
(Spanish
version)
Age at outcome:
12-24 mo
Maternal IQ (WAIS),
maternal age, infant
weight, length, SES,
infant age and sex,
current infant BLL
5.27 (1.93) pg/dL
4.74 (1.96) pg/dL
4.98 (1.93) pg/dL
A large number of results were
obtained from the mediation
analysis. In summary, T1, T2,
and T3 BLLs were associated
with nonsignificant decreases in
12-month PDI. This association
persisted for 24-month PDI at
less magnitude for T2 Pb.
Beta for 12-month PDI
Infant blood GM (SD):
12 mo: 3.92 (1.80) pg/dL
24 mo: 3.49 (1.93) pg/dL
T1
T2
T3
-0.24 (-0.95, 0.48)
-0.38 (-1.10, 0.35)
-0.33 (-1.06, 0.40)
Beta for mediation by GCNT1
cg18515027 methylation of In-
transformed T2 BLLs and 12-
month PDI:
Indirect: 1.25 (-0.11, 3.32)
3-412
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tShekhawat et n: 117
al. (2021)
Western
Rajasthan
India
2018-2019
(enrollment)
Follow-up at
6.5 mo
(average)
Cohort
Mother-child
pairs in third
trimester or at
delivery
Blood
Cord blood; ICP-OES
GM = 4.14 |jg/dL; mean = 4.77 ± 3.3 pg/dL;
median = 4.23 pg/dL
75th: 5.1 pg/dL
Motor
development
assessed using
BSID-III
Age at outcome:
6.5 mo
Maternal age,
gravida, gestational
age, maternal
education, child sex
and weight, preterm
birth, maternal food
intake during
pregnancy, smoking,
alcohol consumption,
maternal residential
and occupational
history, delivery type
(3 (95 % CI)
Umbilical cord Pb level <5
pg/dL (n = 70)
Composite motor:
-0.048 (-0.28, 0.19)
Subscale fine motor:
-0.10 (-1.80, 0.68)
Subscale gross motor:
0.14 (-0.84, 0.94)
Umbilical cord Pb level 5.0-
10.5 pg/dL (n = 47)
Composite motor:
0.01 (-1.19, 0.23)
Subscale fine motor:
0.03 (-3.34, 4.1)
Subscale gross motor:
-0.29 (-5.00, 0.11)
Henn et al.
(2012)
Mexico City
Mexico
1997-2000
(enrollment)
Followed for
24 mo
N: 455
Blood
Women Child venous blood, ICP-MS
recruited during
pregnancy or at
delivery
Age at measurement: 12, 24 mo
12 mo mean (SD): 5.1 (2.6) pg/dL
24 mo mean (SD): 5.0 (2.9) pg/dL
PDI assessed
using BSID-II
(Spanish
version)
Age at Outcome:
12, 18, 24, 30,
36 mo
Sex, gestational age,
hemoglobin, maternal
IQ, maternal
education, and visit
Beta
12-month BLL:
0.02)
24-month BLL:
0.17)
-0.27 (-0.56,
-0.18 (-0.53,
Cohort
3-413
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tParaiuli et al.
Pregnant
Blood
PDI assessed
Maternal age and
Beta
(2015a)
women visiting
using BSID-II
education, BMI,
-4.83 (-16.53, 6.86)
the Bharatpur
Cord blood; ICP-MS, measured for Pb, As and
gestational age,
Chitwan,
General
Zn
Age at outcome:
family income, parity,
Bharatpur
Hospital
24 mo
birth weight, weight at
District
n: 100
Age at measurement:
24 mo, child age
Nepal
Birth cohort:
At birth
assessment, As, Zn,
HOME score
Sep-Oct 2008
women were
Median: 2.06 |jg/dL
(smoking and alcohol
Followed
selected if living
Max: 22.08 pg/dL
consumption not
through 24 mo
in the study
included given low
area for at least
prevalence)
Cohort
2 yr and were at
term pregnancy
(>37 wk of
gestation)
tParaiuli et al.
Birth cohort
Blood
PDI assessed
Maternal age and
Beta
(2015b)
from Bharatpur
using BSID-II
education, BMI,
-2.56 (-9.71, 4.59)
General
Cord blood; ICP-MS, measured for Pb, As and
gestational age,
Chitwan,
Hospital
Zn
Age at outcome:
family income, parity,
Bharatpur
n: 100
36 mo
birth weight, weight at
district
Age at measurement:
24 mo, child age at
Nepal
Resided in area
for at least 2 yr
At birth
assessment, As, Zn,
HOME score
Sep-Oct 2008
delivered at
Median: 2.06 pg/dL
(smoking and alcohol
Followed
term (i.e., >37
Max: 22.08 pg/dL
consumption not
through 36 mo
wk)
included given low
given low prevalence)
Cohort
3-414
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tJianq et al.
(2022)
Taipei
Taiwan
August 2008-
December
2009
(enrollment)
Follow-up 3 yr
Cohort
N: 53 children
Meconium (n =
36)
Hair (n = 52)
Fingernail (n =
43)
Longitudinal
birth cohort
study at a
medical center
hospital in
northern Taiwan
Meconium, Hair, Fingernail
All metals analyzed using ICP-MS.
Child meconium collected at birth
Child hair and fingernails collected at age 1 mo
Median (min, max):
Meconium: 25.6 (2.00, 8815) ng/g
Hair: 3.61 (0.31, 25.1) pg/g
Fingernail: 0.84 (0.06, 24.3) pg/g
Motor
development
assessed using
BSID-III.
Raw total motor
scores were
standardized to
expected mean
of 100 and SD of
15. Raw fine
motor and gross
motor scores
were
standardized to
expected mean
of 10 and SD of
3.
Age at Outcome:
3 yr
Maternal age and
education, newborn
birth head
circumference and
sex, and As and Cd
levels
Beta for log—10 transformed Pb
and log—10 transformed motor
development score
Meconium
Motor: -0.00001 (-0.021, 0.021)
Fine motor: 0.009 (-0.048,
0.065)e
Gross motor: -0.014 (-0.064,
0.036)e
Hair
Motor: 0.020 (-0.009, 0.049)e
Fine motor: 0.046 (-0.015,
0.107)e
Gross motor: 0.006 (-0.043,
0.054)e
Fingernails
Motor: -0.003 (-0.025, 0.019)e
Fine motor: -0.001 (-0.053,
0.052)e
Gross motor: -0.004 (-0.046,
0.037)e
3-415
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tZhou et al.
Shanghai
(2017)
Stress Birth
Cohort Study
Shanghai
n: 139
China
2010-2012
Women
(enrollment)
enrolled in
Followed for
prenatal clinics
24-36 mo after
of maternity
birth
hospitals during
Cohort
mid-to-late
pregnancy.
tLiu et al.
Guangxi Birth
(2022a)
Cohort Study
N: 703 children
Blood
Maternal blood Pb measured using AAS.
Age at measurement:
28-36 wk of gestation
GM: 3.30 pg/dL
Gross motor and
fine motor
development
assessed using
GDS (Chinese
version)
Age at outcome:
24-36 mo
Maternal age at
enrollment, economic
status, maternal
education, gestational
week, child sex, birth
weight and age
Beta per log—10 transformed BLL
Gross motor development: 3.31
(-6.11, 12.73)c
Fine motor development: 0.49
(-11.27, 12.24)c
Guangxi region
China
July-
September
2015
(enrollment)
Followed until
July-
September
2018 (3 yr)
Pregnant
women
recruited from
eight maternity
and child
healthcare
hospitals in six
cities of
Guangxi
Blood, urine
Prenatal maternal serum (first, second, and third
trimesters)
Infant urine
Age at measurement: NR
Maternal serum med (25th, 75th): 0.78 (0.54,
1.24) pg/L
Gross motor
development
using GDS
(Chinese
version)
Age at outcome:
2.57 (SD: 0.14)
yr
Maternal age, pre-
pregnancy BMI,
children's age,
children's gender,
blood sampling time,
delivery mode,
delivery gestational
week, birth head
circumference.
Beta per In-transformed pg/L
increase in Pb
Overall: -2.321 (-3.614, -1.029)c
Male: -3.426 (-6.162, -0.691 )c
Female: -1.182 (-2.805, 0.442)c
Infant urine med (25th,
pg/L
75th): 0.22 (0.14, 0.37)
Cohort
3-416
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tTavlor et al.
(2015)
Avon
UK
Apr. 1, 1991 —
Dec. 31, 1992
(expected
delivery date)
Followed 10 yr
Cohort
Subsample of
ALSPAC Study
N: 582 child
blood
N:4285
prenatal
maternal blood
Pregnant
women in
former Avon
Health Authority
Blood
Maternal blood collected in early pregnancy
(med: 11 wk of gestation); ICP-MS
Child venous blood
Age at measurement: 30 mo
Mean (SD)
Prenatal: 3.67 (1.47) pg/dL
Child: 4.22 (3.12) pg/dL
Balance (heel-
to-toe test) from
the Movement
Assessment
Battery for
Children
(Movement
ABC)
Age at outcome:
7 yr
Static and
dynamic balance
tests based on
BOTMP
Age at outcome:
10 yr
Sex, passive smoking
at 77 or 103 months
old (weekdays and
weekends), and
concurrent Ca and Fe
intakes
OR for >5 pg/dL vs. <5 pg/dL Pb
Prenatal Pb:
Heel-to-toe test: 1.01 (0.95, 1.01)
Dynamic balance: 1.02 (0.95,
1.09)
Static balance: 0.98 (0.92, 1.06)
Child Pb:
Heel-to-toe test: 0.98 (0.92, 1.05)
Dynamic balance: 1.01 (0.93,
1.09)
Static balance: 1.03 (0.94, 1.12)
3-417
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Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tTavlor et al.
(2018)
Avon
UK
Apr. 1, 1991 —
Dec. 31, 1992
(expected
delivery date)
Followed 10 yr
Cohort
Subsample of
ALSPAC Study
N:1558
Pregnant
women in
former Avon
Health Authority
Blood
Maternal blood; ICP-MS
Age at measurement:
Early pregnancy (med: wk 11 of gestation)
Mean (SD)
Prenatal: 3.66(1.55) pg/dL
Balance (heal-
to-toe test), ball
skills (beanbag
toss), and
manual dexterity
(threading lace
and placing
pegs)
Movement
Assessment
Battery for
Children
(Movement
ABC)
Age at outcome:
7 yr
Sex, maternal
education, smoking in
pregnancy, alcohol in
pregnancy, maternal
age and parity
OR for >5 pg/dL vs. <5 pg/dL Pb
Balance: 0.99 (0.74, 1.33)
Ball skills: 0.88 (0.58, 1.32)
Threading lace: 1.12 (0.83, 1.50)
Peg board - preferred hand: 1.19
(0.88, 1.60)
Peg board - non-preferred hand:
1.14 (0.85, 1.54)
OR for highest quartile (NR) vs.
lowest quartile (<5 pg/dL) of Pb
Balance: 0.98 (0.73, 1.31)
Ball skills: 1.07 (0.71, 1.63)
Threading lace: 1.01 (0.75, 1.35)
Peg board - preferred hand: 1.23
(0.92, 1.66)
Peg board - non-preferred hand:
0.99 (0.73, 1.32)
3-418
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tBoucher et al.
(2016)
Nunavik,
Quebec
Canada
October 2005
- February
2010 (outcome
assessment)
Cohort
Nunavik Child
Development
Study
N: 265 school
children
Phone
recruitment of
children with
umbilical cord
blood samples
obtained under
the Arctic Cord
Blood
Monitoring
Program
Blood
Fine motor
performance on
Santa Ana Form
Board (manual
dexterity), Finger
Tapping (fine
Age at measurement: birth (cord), 11.3 yr (child) motor speed),
and Stanford-
Binet Copying
(visuo-motor
integration)
Age at outcome:
11.3 yr (SD: 0.8)
Cord blood and child concurrent venous blood
ICP-MS
Cord mean, median (SD): 4.7, 3.7 (3.4) |jg/dL
Child blood mean, median, SD: 2.7, 2.0 (2.1)
pg/dL
Form Board: child age
and sex, social
environment,
maternal age, parity,
marital status,
smoking during
pregnancy; Finger
Tapping: child age
and sex, social
environment;
Stanford-Binet: child
age and sex, social
environment, marital
status
Others considered:
adoption status,
primary caregiver's
years of education,
Peabody Picture
Vocabulary Test,
RPM, parity, mother
fluency in
English/French,
assimilation to
Western culture,
alcohol and illicit drug
use during
pregnancy, other
contaminants, nutrient
biomarkers
Beta for log-transformed Pb
Cord blood
Manual dexterity: -0.08df
Fine motor speed: -0.19 (-0.33,
—0.05)bf
Visuo-motor integration: -0.01df
Child blood
Manual dexterity: -0.17 (-0.34,
0)b,f
Fine motor speed: -0.21 (-0.37,
—0.05)bf
Visuo-motor integration: 0.1df
3-419
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tParaiuli et al.
n: 79
(2013)
women living in
Chitwan Valley
the study area
Nepal
(i.e., Chitwan)
Sep-Oct 2008
for at least 2 yr,
Cross-
at term
sectional
pregnancy
when the
mothers visited
the hospital
(more than 37
wk of
gestation), age
of 18-40 yr,
singleton birth,
and no report of
diabetes,
hypertension, or
preeclampsia
tLiu et al.
Birth cohort
(2014d)
n: 415 mother-
child pair (219
Shenzhen,
high Pb group
Guangdong
>4.89 pg/dL at
China
first trimester
Jan 2009-Jan
and 196 low Pb
2010
group <1.96
Followed for
pg/dL)
26-30 wk
Cohort
Pregnant
women
recruited during
the early
pregnancy (10-
14 wk)
Blood
Cord blood Pb concentrations determined using
ICP-MS.
Age at measurement:
delivery
mean: 31.7 |jg/L; median: 20.6 |jg/L
75th: 35.1 pg/L
Max: 220.8 |jg/L
Neurodevelopm
ent assessed
using Brazelton
NBAS III
Age at outcome:
1 day old
Maternal age, parity,
mother's education
level; annual family
income, mother's
BMI, birth weight,
gestational age, age
of baby at NBAS
assessment
13 (95 % CI) change in score per
1 pg/L increase in blood Pb
Habituation: 1.44 (-1.19, 4.07)
Orientation: -0.12 (-8.34, 8.10)
Motor system: -2.15 (-4.27,
-0.03)
State organization: 2.15 (-1.58,
5.88)
State regulation: -0.75 (-3.86,
2.36)
Autonomic Stability: 0.71 (-0.48,
1.90)
Abnormal reflex: 1.07 (-1.32,
3.46)
Blood
Maternal, cord blood analyzed using GFAAS.
Maternal BLL classified as low or high
Age at measurement:
First, second and third trimester and at delivery
Low and High BLL groups: First trimester: 1.22
pg/dL and 6.49 pg/dL; second trimester: 1.01
pg/dL and 5.63 pg/dL; third trimester: 1.19 pg/dL
and 6.31 pg/dL; and delivery: 1.26 pg/dL and
6.65 pg/dL
Neurodevelopm
ent assessed
using NBNA
Age at outcome:
3 days
Infant sex, maternal
hemoglobin, IQ,
tobacco use and
parents' occupation,
education, yearly
household income.
Beta for change in NBNA score
per log-transformed Pb
T1
T2
T3
-4.86 (-8.831, —0.889)f
-3.98 (-8.180, 0.220)f
-3.65 (-6.609, 1.309)f
Cord: -3.39 (-7.531, 0.751)
3-420
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Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tNozadi et al.
(2021)
Navajo Nation
United States
Enrollment
February 2013
- June 2018
Follow-up at
age 10-13 mo
Navajo Birth
Cohort Study
(NBCS)
n: 327
Blood
Maternal blood Pb from the 36-week visit or at
the time of delivery was processed using ICP-
DRC-MS.
Age at Measurement:
Mean (SD) maternal age at birth = 27.4 (5.87)
years. Children assessed at 10 and 13 mo.
GM = 0.410 |jg/dL; median
75th: 0.51 pg/dL
95th: 1.20 pg/dL
0.37 pg/dL
Neurodevelopm Multivariable linear Beta
ent assessed
using Ages and
Stages
Questionnaire
Inventory
(ASQ:I)
Age at outcome:
10, 13 mo
regression for fine
motor adjusted for
blood cadmium, urine
cesium, urine arsenic,
and mother's
education; gross
motor adjusted for
urine strontium
Fine motor: -0.63 (-1.19, -0.08)
Gross motor: 0.14 (-0.47, 0.75)
Cohort
tKao et al.
(2021)
Taipei
Taiwan
2011-2014
Cross-
Sectional
recruited from
Taipei MacKay
Memorial
Hospital
n:139 children
less than 3 yr of
age
Hair, fingernails
Pb concentrations in hair and fingernails were
measured using ICP-MS
Age at Measurement:
Mean (SD) 2.8 (0.4) years (children under 3 yr)
GM (SD): hair 2.9 (4.8) i1/4g/g, nails 0.8 (5.1)
TVig/g
Motor
development
assessed using
BSID-III
Age at outcome:
2.8 ± 0.4 yr
Sex, gestational age
at birth, age of the
house (years), leafy-
vegetable intake
(servings/week), and
the area of surface
roads within 100 m of
the residence
Regression results were not
reported because they were not
statistically significant.
3-421
-------
Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tNvanza et al.
(2021)
Northern
Tanzania
Tanzania
2015-2017
(enrollment)
Followed for
12 mo
Cohort
Mining and
Health
Prospective
Longitudinal
Study in
Northern
Tanzania
n: 439
Birth cohort of
mother-child
pairs recruited
in 2nd trimester
Maternal dried blood spots; ICP-MS, measured
for Pb, Hg, and Cd
Age at measurement:
T2
Median: 2.72 |jg/dL
75th: 4.25 pg/dL
Max: 14.5 pg/dL
Gross motor and
fine motor
development
assessed using
MDAT.
Scores in each
domain
classified as
normal (>90th
percentile on all
items in that
domain or <90th
percentile on
one or two items
in the domain) or
impaired (<90th
percentile on
more than two
items in a
domain).
Maternal age and
education, maternal
and paternal
occupation, number
siblings under 5 yr at
home, and family
SES, infant sex, age,
birth weight, height
and weight as a proxy
for nutritional status.
(Covariates with p <
0.20 retained in the
final models.)
Prevalence ratio
Gross motor development: 1.0
(0.9, 1.0)
Fine motor development: 1.0
(0.9, 1.0)
Age at outcome:
between 6 and
12 mo
3-422
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Reference
and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
tPalaniappan n = 755 school Blood
et al. (2011) children (age 3-
7 yr)
Chennai
India
2003-3006
Cross-
sectional
Child venous blood; LeadCare Analyzer
Children . x
attending public A9e at measurement: 3-7 yr
schools in
Chennai Mean (SD): 11.5 (5.3) pg/dL
(kindergarten -
1st grade)
Visual-motor
(drawing),
visual-spatial
(matching), fine
motor
(pegboard)
subtests and
composite
assessed using
WRAVMA
Age at outcome:
3-7 yr
Gender, age,
hemoglobin level,
average monthly
income of the family
(categorical) and
parent education
(categorical)
Beta
Drawing: -0.29 (-0.51, -0.07)
Matching: -0.14 (-0.31, 0.02)
Pegboard: -0.19 (-0.38, 0.01)
Composite: -0.26 (-0.45, -0.07)
Standardized
scores (mean:
100, SD: 15)
AAS = atomic absorption spectrometry; As = arsenic; BASC = Behavior Assessment System for Children; BDI = Beck Depression Inventory; BLL = blood lead level; BOTMP =
Bruininks-Oseretsky Test of Motor Proficiency; BPA = bisphenol A; BRIEF = Behavior Rating Inventory of Executive Functions; BSID = Bayley Scales of Infant and Toddler
Development; CBCL = Child Behavior Check List; CCAAPS = Cincinnati Childhood Allergy and Air Pollution Study; CHECK = Health and Environmental Chemicals in Korea; CI =
confidence interval; CKD = chronic kidney disease; CKiD = Chronic Kidney Disease in Children; CPRS = Conners' Parent Rating Scale; C-TRF = Caregiver-Teacher Report Form;
CTRS = Conners' Teacher Rating Scale; DSM = Diagnostic and Statistical Manual of Mental Disorders; ELEMENT = Early Life Exposure in Mexico to Environmental Toxicants; FSIQ
= full-scale intelligence quotient; GDS = Gesell Development Schedules; GFAAS = graphite furnace atomic absorption spectrometry; GM = geometric mean; Hgb = hemoglobin;
HOME = Health Outcomes and Measures of the Environment; IQ = intelligence quotient; MOCEH = Mothers' and Children's Environmental Health; NBAS = Neonatal Behavioral
Assessment Scales; NBNA = Neonatal Behavioral Neurological Assessment; NHBCS = New Hampshire Birth Cohort Study; NR = not reported; OR = odds ratio; Pb = lead; PDI =
Psychomotor Developmental Index; PHDCN = Project on Human Development in Chicago Neighborhoods; SDQ = Strengths and Difficulties Questionnaire; SES = socioeconomic
status; SRP = self-report of personality; SRS = Social Responsiveness Scale; T1 = first trimester of pregnancy; T2 = second trimester of pregnancy; T3 = third trimester of
pregnancy; WAIS = Weschler Adult Intelligence Scale; WISC-R = Wechsler Intelligence Scale for Children; WRAVMA = Wide Range Assessment of Visual-Motor Abilities; Zn = zinc.
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
The CI was calculated from a p-value and the true CI may be wider or narrower than calculated.
°Results are unstandardized because the Pb level distribution data was not available.
dResults are unstandardized because they did not have an associated SE, CI, or p-value reported in the study.
eResults are unstandardized because the biomarker used for Pb exposure measurement is not blood, tooth, or bone.
'Results are unstandardized because the log base used for exposure transformation was unspecified in the study.
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-423
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Table 3-11T
Animal toxicological studies of Pb exposure and motor function
Study
Species ^Stock/Strain), Timing o, Exposure BLL as Reported (pg/dL)
Endpoints Examined
Flores-Montova and Sobin
(2015)
Mouse (C57BL/6)
Control (distilled water),
M/F, n = 19 (8/11)
30 ppm, M/F, n = 26
(16/10)
230 ppm, M/F, n = 16
(12/4)
PND 0 to PND 28 Oral
Oral,
PND 28 - Males:
drinking
water
0.2 |jg/dL for Control
Oral,
lactation
3.93 |jg/dL for 30 ppm
9.39 |jg/dL for 230 ppm
PND 28 - Females:
0.19 |jg/dL for Control
3.19 |jg/dL for 30 ppm
12.14 |jg/dLfor230 ppm
Oral,
8 wk:
drinking
water
1.8 |jg/dL for Control
21.7 |jg/dL for 250 mg/L
PND 28: OFT, Rotarod Test
Zou et al. (2015)
Mouse (ICR)
Control (distilled water),
M, n = 10
250 mg/L solution, M, n
= 10
-5 wk to 8 wk
8 wk: Rotarod Test,
Locomotor Activity
Rao Barkur and Bairv (2016)
Rat (Wistar) PG: GD -30 to GD Oral,
Control (tap water), M, n 0 lactation
= 12 In utero
G: GD 1 to GD 21
0.2% solution, PG, M, n
= 12 L: PND 1 to PND
21
0.2% solution, G, M, n =
12
0.2% solution, L, M, n =
12
PND 22:
0.19 |jg/dL for Control
3.03 |jg/dL for PG
5.51 |jg/dL for G
26.86 |jg/dL for L
PND 3, 4, 5: Surface Righting
Reflex, PND 6, 8, 10, 12:
Swimming Performance, PND
8, 10, 12: Negative Geotaxis,
PND 14-18: Ascending Wire
Mesh,
3-424
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Betharia and Maher (2012)
Rat (Sprague Dawley)
PND 24:
Control (RO Dl water),
M/F, n = 11-13
GD 0 to PND 20
Oral,
lactation
In utero
PND 2:
1.77 ng/g (0.188 pg/dL) for
Control
85.17 ng/g (9.02 pg/dL) for 10
PND 1-10: Surface Righting
Reflex, PND 24, 59: OFT
10 |jg/mL, M/F, n = 11-
13
|jg/mL
PND 25:
PND 59:
0.83 ng/g (0.088 pg/dL) for
Control (RO Dl water),
Control
M/F, n = 10-11
9.21 ng/g (0.98 pg/dL) for 10
10 |jg/mL, M/F, n = 10-
11
pg/mL
PND 60:
0.23 ng/g (0.024 pg/dL) for
Control
0.30 ng/g (0.032 pg/dL) for 10
pg/mL
3-425
-------
Study
Species (Stock/Strain), Timing of
n, Sex Exposure
EDe°aSi|Js0 BLL as ReP°rted (M9^L)
Endpoints Examined
Basha and Reddv (2015)
Rat (Wistar)
Control (deionized
water), M, n = 8
0.2 % solution, M, n = i
GD 6 to GD 21 Inutero PND21:
0.21 |jg/dL for Control
11.2 |jg/dL for 0.2% solution
PND 28:
0.33 |jg/dL for Control
12.3 |jg/dL for 0.2% solution
4 mo:
0.19 |jg/dL for Control
5.9 |jg/dL for 0.2% solution
PND 4-7: Surface Righting
Reflex, PND 8-10: Negative
Geotaxis, PND 12-16:
Forelimb Hang, PND 21, PND
28, 4 mo: Locomotor Activity
Tartaalione et al. (2020)
Rat (Wistar) GD -28 to PND 23
Oral,
PND 23:
PND 4, 7, 10, 12: Neonatal
Control (tap water), M/F
lactation
Spontaneous Movement,
n = 16 (9/7)
In utero
0.007 |jg/mL (0.7 pg/dL) for
Control
PND 4, 7, 10, 12: Surface
Righting Reflex, PND 4, 7, 10,
50 mg/L, M/F, n = 16
12: Negative Geotaxis, PND
(9/7)
0.255 pg/mL (25.5 pg/dL) for 50
mg/L
30: OFT
Faulk et al. (2014)
Mouse (Agouti)
GD-14 to PND 21 Oral,
PND 21 (Maternal BLL):
PND 90, 180, and 270:
Control (distilled water),
lactation
Locomotor Activity
M/F, n = 30
In utero
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Bashaetal. (2014)
Rat (Not Specified)
PND 1 to PND 21
Oral,
PND 45:
PND 45, 4 mo, 12 mo, 18 mo:
Control (deionized
lactation
OFT, Locomotor Activity
water), M, n = 6
0.42 pg/dL for Control
0.2% solution, M, n = 6
49.5 pg/dL for 0.2% solution
4 mo:
0.56 pg/dL for Control
14.4 pg/dL for 0.2% solution
12 mo:
0.46 pg/dL for Control
6.96 pg/dL for 0.2% solution
18 mo:
0.12 pg/dL for Control
11.2 pg/dL for 0.2% solution
Mansouri et al. (2012)
Rat (Wistar)
Control (distilled water),
M/F, n = 16 (8/8)
50 mg/L, M/F, n = 16
(8/8)
PND 70 to PND
100
Oral,
drinking
water
PND 100 - Males:
2.05 pg/dL for Control
8.8 pg/dL for 50 mg/L
PND 100 - Females:
PND 100: OFT, Rotarod Test
2.17 |jg/dL for Control
6.8 |jg/dL for 50 mg/L
3-427
-------
Study
Species (Stock/Strain), Timingof Exposurc BLL as Reported (Mg/dL) Endpoints Examined
Duan et al. (2017) Mouse (CD1) PND 1 to PND 21 Oral, PND21: PND 7, 11, 15, 19: OFT
Control (distilled water), lactation
M/F, n = 5 16.2 pg/L (1.6 pg/dL) for Control
27 ppm, M/F, n = 5 191.8 pg/L (19.2 pg/dL) for 27
ppm
109 ppm, M/F, n = 5
283.4 pg/L (28.3 pg/dL) for 109
ppm
PND 35:
14.3 pg/L (1.4 pg/dL) for Control
283.4 pg/L (28.3 pg/dL) for 27
ppm
376.9 pg/L (37.7 pg/dL) for 109
ppm
Wana et al. (2016)
Rat (Sprague Dawley)
PND 24 to PND 56
Oral,
PND 56: PND 60-66: OFT
Control (tap water), M, n
drinking
= 7
water
11 pg/L (1.1 pg/dL) for Control
100 ppm, M, n = 9
133 pg/L (13.3 pg/dL) for 100 ppm
Shvachiv et al. (2018)
Rat (Wistar)
Intermittent
Oral,
PND 196: PND 189: OFT
Control (tap water), M/F,
Exposure: GD 7 to
drinking
n = 8
PND 84, PND 140
water
<0.1 pg/dL for Control
to PND 196
Oral,
0.2% (p/v) solution
lactation
18.8 pg/dL for 0.2% (Intermittent)
(distilled water), M/F, n =
Continuous
In utero
9 - Intermittent
Exposure: GD 7 to
24.4 pg/dL for 0.2% (Continuous)
exposure PND 196
0.2% (p/v) solution, M/F,
n = 9 - Continuous
exposure
3-428
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Stansfield et al. (2015)
Rat(Long-Evans)
Control (chow), M/F, n =
11-23
1500 ppm, M/F, n = 11-
23
GD Oto PND 50
Oral, diet
Oral,
lactation
In utero
PND 50:
0.6 pg/dL for Control
22.2 pg/dL for 1500 ppm
PND 50: Locomotor Activity
Neuwirth et al. (2019a)
Rat(Long-Evans)
Control (tap water), M/F,
n = 48 (30/18)
150 ppm, M/F, n = 62
(32/30)
1000 ppm, M/F, n = 49
(30/19)
GD 0 to PND 22
Oral,
lactation
In utero
PND 22:
-------
Study
Species (Stock/Strain), Timingof Exposurc BLL as Reported (Mg/dL) Endpoints Examined
Sobolewski et al. (2020) Mouse (C57BL/6)
F0:
Control (distilled Dl
water), F, n = 10
100 ppm, F, n = 10
F1:
see Figure 1, n = 12
F2:
see Figure 1, n = 12
F3:
see Figure 1, n = 8-10
Sinah et al. (2019)
Rat (Wistar) 3 mo to 6 mo
Control (distilled water),
M, n = 5
2.5 mg/kg, M, n = 5
Oral,
gavage
6 mo:
5.76 |jg/dL for Control
28.4 |jg/dL for 2.5 mg/kg
6 mo: Locomotor Activity,
Rotarod Test
Viaueras-Villasenor et al.
Rat (Wistar) GD 0 to PND 21
Oral,
PND 110:
PND 90 to PND 110:
(2021)
Control (tap water), M, n
lactation
Locomotor Activity
= 8
In utero
2.04 |jg/dL for Control
320 ppm, M, n = 8
26.3 |jg/dL for 320 ppm
F1: GD -60 to PND Oral,
23-27 lactation
In utero
F1 PND 6-7:
0 |jg/dL for Control
12.5 |jg/dL for 100 ppm (F0
dosing)
F3 PND 6-7:
0 ng/dL for Control
0 |jg/dL for 100 ppm (F0 dosing)
PND 60-120 (variable by
endpoint): Locomotor Activity
3-430
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Al-Qahtani et al. (2022)
Mouse (Albino)
Control (distilled water),
M, n = 10
0.2 mg/kg, M, n = 10
8-9 wk to 14-15 wk
Oral,
gavage
14-15 wk:
1.2 |jg/100 mL (1.2 pg/dL) for
Control
7.1 |jg/100 mL (7.1 pg/dL) for 0.2
mg/kg
NR: Locomotor Activity
BLL = blood lead level; F# = filial generation; F = female; GD = gestational day; LOD = limit of detection; M = male; MRI = magnetic resonance imaging; mo = month(s); NaAc =
sodium acetate; NR = not reported; OFT = open-field test Pb = lead; PG = pregestation; PND = postnatal day; wk = week(s); yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
3-431
-------
Table 3-12E Epidemiologic studies of Pb exposure and sensory organ function in children
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% CIs
Dietrich et al. (1992)
Cincinnati, U.S.
Cross-sectional
The Cincinnati
lead study cohort
n: 259
Blood
Age at measurement: prenatal-
yr
Mean (SD) |jg/dL:
Prenatal 8.2 (3.8)
Neonatal 4.8 (3.3)
Central auditory
processing abilities and
cognitive developmental
status
Age at outcome: 5 yr
Measures of fetal distress
and growth, perinatal
complications, postnatal
indices of health and
nutritional status,
sociodemographic
characteristics, and
psychosocial features of the
home environment
Betab
Filtered Word
Score (total
number of words
correctly identified
in both ears)
Prenatal: -0.12
Neonatal: -0.26
Mean lifetime
through 5 yr:
-0.07
Schwartz and Otto (1991)
HHANES, U.S.
Cross-sectional
Hispanic Health Blood
and Nutrition
Examination
Survey
n: 3545
Age at measurement: 6-19 yr
Median (25th' 75th) pg/dL:
Mexican Americans 8 (6, 11)
Cuban American 8 (6, 10)
Puerto Ricans 8 (6, 11)
Elevated hearing
threshold
Audiometric evaluations
were performed for all
subjects Beltone model
200-C audiometers were
used in the survey;
Hearing threshold was
defined as the lowest
intensity of a pure tone
that was just audible to
the subject.
NR
An increase15 in
BLL from 7
~microg/dl to 18
pg/dl was
associated with an
approximately 2-dB
loss of hearing at
all frequencies
Age at outcome: 6-19 yr
3-432
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% CIs
Schwartz and Otto (1987)
NHANES II
U.S.
Cross-sectional
NHANES II Blood
n: 4519 Age at measurement: 4-19 yr
Range of Pb: 6 to 47 |jg/dL
Hearing thresholds
Standard Beltone Model
200C audiometers were
used and ca liberated
weekly with B&K Model
2203 sound level meters
in accordance with 1969
ANSI specifications.
Tests were conducted at
500, 1000, 2000, and
4000Hz on each ear.
Race, lead, ear discharge,
cold in last 2-week, other
ear condition, chronic ear
discharge, income, dietary
calcium, sex, current cold,
ringing in ear(s), earache,
previous running ear,
diagnosed hearing
impairment, degree of
urbanization, head of
household education level
The risk of
elevated hearing
thresholds at 500,
1000, 2000 and
4000 Hz increased
with increasing
PbB for both ears
Age at outcome: 4-19 yr
tYin etal. (2021)
n: 234-7596 in 8 Blood
studies
Age at measurement: 3-87 yr
Hearing loss
All studies included in the
meta-analysis controlled for
age and sex. Adjustment for
other potential confounders
varies by studies, but
includes monthly income,
education levels, smoking
status, BMI, ethnicity, work
duration, ototoxic
medication, blood lead,
occupational noise, loud
noise, and firearm noise,
and hypertension and
diabetes
OR (95% Cl)b
1.53 (1.24,1.87)
tChoi and Park (2017)
Korea
2010-2012
KNHANES
n: 5187 adults
and 853
adolescents
Blood
Hearing loss (>15dB) at
speech frequency;
Hearing loss (>15dB) at
high frequency
Age, age squared, sex,
education, BMI, current
cigarette smoking
OR (95% Cl)b
Hearing Loss (>15
dB) High-frequency
PTA
Pb Quartile 2
(0.978-1.260):
0.89 (0.39, 2.03)
3-433
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% CIs
Cross-sectional
Graphite furnace atomic
absorption spectrometry
Age at Measurement:
adolescents 12-19 yr (mean±SE
15.6 ± 0.10)
Geometric mean (95% CI) (age-
adjusted): 1.26 [jg/dL (1.22, 1.30)
Pure-tone air conduction
hearing thresholds were
obtained for each ear at
frequencies of 0.5, 1, 2, 3,
4, and 6 kHz over an
intensity range of -10
to 110 dB -10 to 110 dB.
Age at outcome: 12-19 yr
Pb Quartile 3
(1.261-1.557):
1.88 (0.83, 4.25)
Pb Quartile 4
(1.562-5.904):
1.38 (0.63, 3.02)
Per doubling of Pb:
1.26 (0.73, 2.16)
Hearing Loss (>15
dB) Speech-
frequency PTA
Pb Quartile 2
(0.978-1.260):
1.17 (0.41, 3.32)
Pb Quartile 3
(1.261-1.557):
1.08 (0.38, 3.08)
Pb Quartile 4
(1.562-5.904):
1.24 (0.34, 4.49)
Per doubling of Pb:
1.2 (0.48, 3.05)
3-434
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% CIs
tXu et al. (2020)
China
October-December 2014
Cross-sectional
n: 116
Blood
Graphite furnace atomic
absorption spectrometry (GFAAS,
Jena Zeenit 650, Germany)
Age at measurement: 3-7 yr
Median ± SEM (P25, P75):
Exposed group
5.29 ± 0.29 (3.61, 7.40)
Refence group
3.63 ± 0.24 (2.98, 4.77)
DNA methylation and
hearing loss
Age at outcome: 3-7 yr
Both continuous variables
for child age, gender,
weight, height and BMI, and
categorical variables for
presence of family member
smoking, residence distance
to the road, residence
nearby noise, residence
renovation noise within a
year, often listening music
with earphones within a
year, often watching
television programs in loud
noise, and often play (i.e.,
toys or music, etc.) in loud
noise
Beta (95% Cl)b
Q1 0.139 (0.007,
2.968
Q2 0.051 (0.003,
0.977)
Q3 0.16 (0.016,
1.58)
Q4 2.765 (1.795,
15.237)
OR (95% CI)
Hearing loss in
both ears 1.40
(1.06, 1.84)
Left ear 1.46 (1.12,
1.91)
tSharaorodskv et al. (2011) NHANES
n: 2535
NHANES, U.S.
2005-2008
Cross-sectional
Blood
Inductively coupled plasma mass
spectrometry
Age at Measurement: 12-19 yr
Weighted Mean (95% CI):
Age 12-13: 1.00 pg/dL
(0.92-1.09 pg/dL)
Age 14-15: 0.93 pg/dL
(0.87-0.99 pg/dL)
Age 16-17: 0.85 pg/dL
(0.79-0.91 pg/dL)
Age 18-19: 0.93 pg/dL
(0.84-1.03 pg/dL)
Any Hearing Loss (>15
dB),
High-Frequency Hearing
Loss, Low-Frequency
Hearing Loss
Age at outcome: 12-19 yr
Age, sex, race-ethnicity,
PIR, history of 3 or more ear
infections, loud noise
exposure, and smoking
OR (95% Cl)b (<1
pg/dL reference)
Any >15 dB
1-1.99 pg/dL 0.99
(0.67-1.46)
>2 pg/dL 1.95
(1.24-3.07)
High-Frequency
1-1.99 pg/dL 1.20
(0.80-1.80)
>2 pg/dL 2.22
(1.39-3.56)
Low-Frequency
1-1.99 pg/dL 1.24
(0.82-1.86)
>2 pg/dL 1.13
(0.61-2.07)
3-435
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% CIs
tLiu etal. (2018c)
Guiyu (e-waste recycling
area) & Haojing (exposure
control, no e-waste
processing), China
2014
Cross-sectional
n: 234 (146
exposed; 88
reference)
Blood
Graphite furnace atomic
absorption spectrometry (GFAAS,
Jena Zeenit 650, Germany)
Age at Measurement:
3-7 yr
Mediant SE: 4.94 ± 0.20 |jg/dL in
exposed; 3.85 ± 1.81 pg/dL in
reference
Hearing loss, Low
frequency hearing loss,
High frequency hearing
loss
Age at outcome: 3-7 yr
Child age, gender, weight,
height, BMI, parent
education level, family
member smoking, family
monthly income, residence
distance to the road,
residence nearby noise,
residence renovation noise
within a year, often listening
to music with earphones
within a year, often watching
television programs in loud
noise, and often play (i.e.,
toys or music, etc.) in loud
noise
OR (95% Cl)b
Hearing loss total
1.24 (1.029, 1.486)
Low frequency
1.02 (0.869, 1.190)
High frequency
1.08 (0.839, 1.379)
3-436
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% CIs
tPawlas et al. (2015)
Upper Silesia, Poland
1996-2001 and 2008-2010
Cross-sectional
Two cohorts
merged
n: 483
Blood
Graphite furnace atomic
absorption spectrometry
Age at Measurement:
4-13 yr
Median: 4.50 |jg/dL
Pure-tone audiometry
(PTA),
Brainstem auditory
evoked potentials
(BAEP),
Acoustic otoemission
Age at Outcome: 4-13 yr
Cohort, mother's education
(dichotomized into
'secondary school or
higher', or 'less',
corresponding to primary
and apprenticeship) and
smoking during pregnancy,
and the child's sex, birth
weight, apgar score, history
of mumps, age, and
pressure in middle ear on
both sides
Beta (95% Cl)b
ALAD MspI
ALAD1-1 0.3
(0.15, 0.45)
ALAD*2 0.42
(-0.03, 0.87)
ALAD Rsa1
TT+TC 0.3 (0.1,
0.5)
CC 0.2 (-0.05,
0.45)
VDR Bsml
bb 0.03 (-0.22,
0.28)
Bb+BB 0.4 (0.25,
0.55)
VDR taq1
TT 0.04 (-0.21,
0.29)
Tt+tt 0.4 (0.2, 0.6)
VDR fokl
FF+Ff 0.4 (0.25,
0.55)
ff -0.1 (-0.6, 0.4)
3-437
-------
tSilveret al. (2016)
Sanhe County, Hebei
Province, China
November 2009- November
2011
Cohort
n: 391 (ARB:
auditory
brainstem
response), 1148
(VA: visual
acuity)
Maternal Blood
AAS
Age at Measurement:
Pregnant woman 18 yr or older
Mean (SD) gestational age at
mid-pregnancy visit
ABR subset 15.7 (2.2) weeks
VA subset 15.5 (1.9) weeks
Mean (SD) gestational age at
late-pregnancy
ABR subset 38.8 (1.3) weeks
VA subset 39.3 (1.3) weeks
Mean (SD) gestational age at
birth
ABR subset 39.2 (1.1) weeks
VA subset 39.7 (1.1) weeks
ABR Pb median
2.9 |jg/dL at mid-pregnancy
3.0 |jg/dL at late-pregnancy
<2.0 |jg/dL at birth (cord blood)
GM (SD)
2.4 (2.5) |jg/dL at mid-pregnancy
2.7 (2.3) |jg/dL at late-pregnancy
<2.0 |jg/dL at birth (cord blood);
VA median
2.9 |jg/dL at mid-pregnancy, 3.3
|jg/dL at late-pregnancy, 2.1
|jg/dL at birth (cord blood)
GM (SD)
2.4 (2.6) |jg/dL at mid-pregnancy
2.9 (2.2) |jg/dL at late-pregnancy
<2.0 |jg/dL at birth (cord blood)
3-438
ABR;
Grating visual acuity (VA)
Age at Outcome:
ABR mean 2 d old
VA mean 6 wk old
Sex, age attesting, cord
blood iron status,
gestational age, birth
weight, head circumference
Mid pregnancy
lead Med. (2-3.8
|jg/dL) 0.02 (-0.01
-0.05)
Late-pregnancy
lead High (>3.8
|jg/dL) 0.05 (0.02,
0.08)
Late-pregnancy
lead Med. (2-3.8
|jg/dL) 0.03 (0.01,
0.06)
Cord lead High
(>3.2 [jg/dL) 0
(-0.02, 0.03)
Cord lead Med. (2-
3.2 [jg/dL) 0
(-0.03, 0.03)
Beta (95% Cl)b
ARB C-P ratio
Mid pregnancy
lead High (>3.8
|jg/dL) 0.02 (-0.01
-0.05)
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% CIs
tAlvarenqa et al. (2015)
Brazil
Followed 35.5 mo
Contemporary cross-sectional
cohort
n: 130 children
(80 males & 50
females)
Blood
AAS with graphite furnace
Age at Measurement:
18 mo-14 yr (Mean: 6 yr 8 mo ±
2 yr 3 mo)
Mean: 12.2 pg/dL; SD = 5.7
pg/dL
Median: 10.2 pg/dL
Auditory brainstem
response
Age, gender, cumulative
blood lead levels, and
date of the audiological
Age at outcome: 18 mo- assessment
14 yr
Beta (95% Cl)b
Wave III, in relation
to wave I
Constant 4.00
(3.97, 4.04)
Wave I RE 0.58
(0.44, 0.72)
Male RE 0.09
(0.05, 0.13)
Constant 4.03
(3.99, 4.06)
Wave I LE 0.61
(0.45, 0.77)
Male LE 0.07
(0.03, 0.11)
Wave V, in relation
to wave III
Constant 5.77
(5.74, 5.80)
Wave I RE 0.81
(0.68, 0.94)
Male RE 0.073
(0.03, 0.11)
Constant
5.78 (5.75, 5.81)
Wave I LE 0.85
(0.73, 0.97)
Male LE 0.08
(0.05, 0.12)
3-439
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% CIs
tFillion et al. (2013)
n: 228
Blood
Contrast sensitivity
(cycles per degree, cpd);
Age, sex, current smoking
(yes vs. no), current drinking
Beta (95% Cl)b
Spatial frequency
Lower Tapajos River Basin,
Inductively coupled plasma mass
Acquired color vision loss
(color confusion index,
CCI)
(yes vs. no)
with %EPA
1.5 cpd -1.32
(-4.30; 1.65)
State of Para, Brazil
spectrometry (ICP-MS, Perkin
Elmer DRC II)
May to July 2006
Age at Measurement:
Age at outcome: 15-66 yr
3 cpd 2.06 (-2.87;
6.99)
15-66 yr (median = 33.0 yr)
6 cpd 0.60 (-6.04;
7.25)
12 cpd -13.33
(-23.28; -3.49)
18 cpd -2.43
(-6.64; 1.79)
CCI 0.16 (-0.03;
0.33)
Cross-sectional
Mean = 12.8 ± 8.4 |jg/dL; Median
= 10.5 |jg/dL
AAS = Atomic absorption spectrometry; ABR = Auditory brainstem response; BLL = blood lead level; CI = confidence interval; OR = odds ratio; Pb = lead; PTA = pure-tone average.
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bEffect estimates are not standardized because data pertaining to the BLL distribution and/or base for the log-transformation were not reported.
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-440
-------
Table 3-13E Epidemiologic studies of Pb exposure, social cognition, and behavior in children
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tKimetal. (2016)
South Korea
2005-2006
(enrollment); 2009-
2010 (follow-up)
Cohort
CHEER study
n: 2,437
Children recruited
from 33 elementary
schools across 10
Korean cities
Blood
Child blood; GFAAS
Age at measurement:
7-8 yr old, 9-10 yr old, and 11-
12 yr old
GM (pg/dL):
7-8 y: 1.64; 9-10 y: 1.58; 11-12
y: 1.58
75th (pg/dL):
7-8 y: 2.36; 9-10 y: 2.08; 11-12
y: 2.05
95th (pg/dL):
7-8 y: 3.47; 9-10 y: 3.05; 11-12
y: 3.05
Autistic behaviors
Parent responses to
ASSQ and SRS
Age at outcome:
11-12 yr
Child sex, fetal and
environmental tobacco
smoke, parental
education levels,
family income, low
birth weight,
breastfeeding,
gestational age, fish
intake, and blood Hg
level
Change in SRS Scores*
Exposure at 7-8 yr
1.37 (0.75, 1.98)
Exposure at 9-10 yr
0.56 (-0.33, 1.44)
Exposure at 11-12 yr
0.39 (-0.47, 1.25)
Change in ASSQ Scores*
Exposure at 7-8 yr
0.09 (0.03, 0.14)
Exposure at 9-10 yr
-0.02 (-0.09, 0.05)
Exposure at 11-12 yr
0.03 (-0.04, 0.10)
*Higher score indicates more
autistic behaviors
OR Autism (ASSQ >17)
Exposure at 7-8 yr
1.45 (1.10, 1.93)
Exposure at 9-10 yr
0.86 (0.60, 1.23)
Exposure at 11-12 yr
0.97 (0.70, 1.35)
3-441
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tArora et al. (2017) Roots of Autism Tooth
Sweden
2011-2016
(enrollment)
Cohort
and ADHD Twin
Study in Sweden
n: 32 twin pairs
and 12 individual
twins
Monozygotic and
dizygotic twins
discordant for
ASD; discordance
defined as >2
points differences
on the Autism-Tics,
ADHD and other
Comorbidities
subscale
Shed deciduous teeth, validated
by maternal, cord, and serial
child blood Pb; laser ablation
ICP-MS
Age at measurement: estimating
various timepoints from 20 wk
prenatal to 30 wk postnatal
Mean NR
ASD diagnosis
ADOS-2, SRS-2
among discordant twins
for ASD (ICD10 [Autism
or Asperger's]; DSM-5
[ASD])
Age at outcome:
8-12 yr
Genetic factors
Child sex, zygosity,
gestational age, the
average birth weight of
the twin pairs, and the
SD of the birth weight
in the twin pairs.
OR of log-transformed Pb for
ASD case vs. non-ASD twin
control: 1.5 (0.9, 2.5)bd
More quantitative results
depicted graphically (see
Figure 3-2)
tSkoaheim et al.
(2021)
Nationwide
Norway
2002-2009
(enrollment)
Case-control
Norwegian Mother,
Father and Child
Cohort Study
(MoBa)
n: 397 ASD cases,
1034 controls
Children from a
birth cohort
Blood
ASD diagnosis
Maternal whole blood; ICP-SFMS NPR
Age at measurement:
wk 17 of gestation
Exposure Quartiles:
Q1: 0.16-0.65 pg/dL
Q2: 0.65-0.86 pg/dL
Q3: 0.86-1.12 pg/dL
Q4: 1.12-8.24 pg/dL
Age at outcome: NR
Birth year and child
sex-matched controls
Child sex, birth weight,
birth year, and SGA,
maternal age at
delivery, education,
parity, pre-pregnancy
BMI, kg/m2), self-
reported smoking and
alcohol intake during
pregnancy, FFQ-
based estimates of
seafood intake (g/day),
and dietary iodine
intake (pg/day)
OR for In-transformed Pb
Q1:
Ref.
Q2:
0.80
(0.57,
1.12)'
Q3:
0.79
(0.56,
1.12)'
Q4:
0.81
(0.57,
1.15)'
3-442
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tRahbar et al. (2015)
Jamaican Autism
Study
Blood
ASD diagnosis
Age, sex-matched
controls
GMD for In-transformed Pb
(ASD Cases vs. Controls):
Kingston
n: 100 cases; 100
Child venous blood; ICP-MS
DSM-IV-TR criteria,
Jamaica
controls
ADOS
maternal age, parental
-0.17 (-0.86, 0.52)bc
Children 2-8 yrat
Age at measurement:
education levels,
December 2009-
2-8 yr
Age at outcome:
parish at child's birth,
March 2012
enrollment
2-8 yr
SES (i.e., car
(enrollment)
GM (SD) (cases): 2.25 (2.23)
pg/dL
ownership by the
family), consumption
Case-control
GM (SD) (controls): 2.73 (1.85)
of shellfish (lobsters,
pg/dL
crabs), and Teflon use
(pots, pans, and
dishes) for cooking
tRahbar et al. (2021)
n: 30 cases; 30
controls
Blood
ASD diagnosis
Age, sex-matched
controls
GMD for In-transformed Pb
(ASD Cases vs. Controls):
Karachi
Child venous blood; ICP-MS
DSM-IV-TR criteria,
Pakistan
children at clinics
ADOS
maternal age, parental
-1.37 |jg/dL (-3.28, 0.54)bc
affiliated with Aga
Age at measurement:
education level, and
Study years NR
Khan University
2-8 yr
Age at outcome:
SES (i.e., car
2-12 yr
ownership by the
Case-control
GM (cases): 7.11 |jg/dL;
GM (controls): 8.48 |jg/dL
family) and dietary
consumptions
dummy variables that
represented the
matched pairs
3-443
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tDonq et al. (2022)
Northeast China
October 2017-
January 2020
(enrollment)
Case-Control
n: 512 children with Blood
ASD
Children diagnosed
with ASD at First
Hospital of Jilin
University
Child serum
Age at measurement: 2-13 yr
Mean (SD)
Mild Autism: 2.58 (1.08) pg/dL
Moderate/severe: 2.58 (1.08)
pg/dL
ASD severity
Severity of autism
symptoms determined
by CARS
Age at outcome 2-13 yr
Age, place of
residence, caregivers,
parental education
level, gastrointestinal
problems.
Also considered sex,
siblings, parental age
at pregnancy,
household income,
family history of mental
illness, vitamin intake
during pregnancy,
eating problems,
sleeping problems,
gastrointestinal
problems, ADHD
comorbidity
Beta
0.03 (0.01, 0.05)c
tRyqiel et al. (2021) ELEMENT project Blood
Orientation/engagement Maternal IQ (WAIS),
Mexico City
Mexico
1997-2005
(enrollment)
Cohort
n: 85
Mother-child pairs
recruited at the
Mexican Social
Security Institute
Maternal and child venous blood;
ICP-MS, GFAAS
Age at measurement:
T1, T2, T3 (maternal); 12, 24 mo
(child)
Maternal blood GM (SD):
T1: 5.27 (1.93) pg/dL
T2: 4.74 (1.96) pg/dL
T3: 4.98 (1.93) pg/dL
and emotional
regulation
ORIEN and EMOCI
scores from BRS of
BSID-IIS
Age at outcome: 12-24
mo
maternal age, infant
weight, length, SES,
infant age and sex,
current infant BLL
A large number of results were
obtained from the mediation
analysis. In summary, T2 BLLs
were consistently inversely
associated with 24-month
EMOCI and ORIEN scores
Beta
24-month EMOCI at T2:
-1.13% (-2.63, 0.37)
24-month ORIEN at T2:
-0.98% (-2.83, 0.88)
Infant blood GM (SD):
12 mo: 3.92 (1.80) pg/dL
24 mo: 3.49 (1.93) pg/dL
3-444
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tShekhawat et al.
(2021)
Western Rajasthan
India
2018-2019
(enrollment)
Followed through 6.5
mo (average)
n: 117
Mother-child pairs
in third trimester or
at delivery
Blood
Cord blood; ICP-OES
Age at measurement:
Delivery
GM = 4.14 |jg/dL; mean = 4.77 ±
3.3 |jg/dL; median = 4.23 |jg/dL
75th: 5.1 |jg/dL
Social-emotional
development score
using BSID-III
Age at outcome: 6.5 mo
Maternal age, gravida,
gestational age,
maternal education,
child sex and weight,
preterm birth, maternal
food intake during
pregnancy, smoking,
alcohol consumption,
maternal residential
and occupational
history, delivery type
13 (95 % CI) for socio-emotional
development scores
Pb < 5 |jg/dL: 0.19 (-0.46,
0.46)
Pb 5.0-10.5 |jg/dL: -0.05
(-0.60, 0.86)
Cohort
tNozadi et al. (2021)
Navajo Nation
United States
February 2013-June
2018 (enrollment)
Followed through 10-
13 mo
Cohort
Navajo Birth
Cohort Study
(NBCS)
n: 327
Children of
mothers (age 14-
45 yr) living across
Navajo Nation with
community
exposure to metal
mixtures from
abandoned
uranium mines
Blood
Maternal blood,
DRC-MS.
child blood; ICP-
Age at measurement:
Delivery or 36-wk visit (maternal);
10, 13 mo (child)
GM = 0.410 |jg/dL;
0.37 |jg/dL
75th: 0.51 pg/dL
95th: 1.20 pg/dL
median =
Communication and
personal-social domain
scores using the ASQ:I.
Age-adjusted scores.
Age at outcome: 10-13
mo
Age.
Also considered
maternal age, marital
status, maternal
occupation and
education, household
income,
concentrations of
various metals in
urine, blood, and
serum
Beta (95% CI)
Communication: -0.15 (-0.58,
0.28)
Personal-Social: -0.11 (-0.72,
0.50)
3-445
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tLin etal. (2013)
Taipei, Taiwan
April 2004-Jan 2005
(enrollment)
Followed through 2 yr
Panel Study
TBPS
n: 230
Singleton full-term
children of non-
smoking mothers
without
occupational
exposure attending
medical center,
hospital, and
clinics in Taipei
Blood
Maternal blood, cord blood; ICP-
MS, measured for Pb, Mn, As,
and Hg.
Pb categories:
Low: <16.45 |jg/L
High: >16.45 pg/L
Mn categories:
Low: <59.59 pg/L
High: >59.59 pg/L
Age at measurement:
Delivery
Mean: 13 pg/L, GM: 10.61 pg/L
75th: 16.45 pg/L
Max: 43.22 pg/L
Social and self-help
ability DQs
CDIIT
Age at outcome:
2 yr
Maternal age,
maternal education,
child sex,
environmental tobacco
smoke during
pregnancy and after
delivery, fish intake,
and HOME Inventory
score
Beta
Social
High vs. Low Pb: -5.89
(-10.81, —0.97)c
High Mn x low Pb: 2.83
(-3.442, 9.102)c
Low Mn x high Pb: -2.9
(-9.231, 3.431 )c
High Mn x high Pb: -7.01
(-14.144, 0.124)c
Self-help
High vs. Low Pb: -1.26
(-5.905, 3.385)c
High Mn x low Pb: 0.49
(-5.429, 6.409)c
Low Mn x high Pb: 0.35
(-5.608, 6.308)c
High Mn x high Pb: -2.38
(-9.103, 4.343)c
tNvanza et al. (2021)
Northern Tanzania
Tanzania
2015-2017
(enrollment)
Followed through 12
mo
Cohort
Mining and Health
Prospective
Longitudinal Study
in Northern
Tanzania
n: 439
Birth cohort of
mother-child pairs
recruited in 2nd
trimester
Maternal dried blood spots; ICP-
MS, measured for Pb, Hg, and
Cd
Age at measurement:
second trimester
Median: 2.72 pg/dL
75th: 4.25 pg/dL
Max: 14.5 pg/dL
Social development
domain using MDAT.
Scores classified as
normal (>90th
percentile on all items in
the domain or <90th
percentile on one or two
items in the domain) or
impaired (<90th
percentile on more than
two items in the
domain).
Age at outcome:
6-12 mo
Maternal age and
education, maternal
and paternal
occupation, number
siblings under 5 yr at
home, and family SES,
infant sex, age, birth
weight, height and
weight as a proxy for
nutritional status
(covariates with p <
0.20 retained in the
final models)
Prevalence ratio:
Social status development:
1.01 (1.00, 1.02)
3-446
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tDohertv et al.
(2020)
New Hampshire
U.S.
2009 to 2014-2019
Cohort
NHBCS
n: 371 (SRS-2);
318 (BASC-2)
Mother-child pairs
Toenails
Maternal and infant toenails;
Median (maternal prenatal): 0.14
|jg/g (SRS), 0.13 pg/g (BASC);
Median (maternal postnatal):
0.10 |jg/g (SRS), 0.11 pg/g
(BASC); Median (infant): 0.35
|jg/g (SRS), 0.37 |jg/g
Composite score
(Social Awareness,
Social Cognition, Social
Communication, Social
Motivation, and
Restricted Interests and
Repetitive Behavior) on
SRS-2.
Adaptive skills
composite on BASC-2;
see also Section 3.5.2.2
Age at outcome:
3 yr old
Maternal age,
maternal BMI, parental
education, maternal
smoking, marital
status, parity, child age
at last breastfeeding,
Healthy Eating Index
score, year of birth,
sex, and age of the
child at testing
Beta per log2-transformed |jg/g
increase in toenail Pb
Total SRS-2
Maternal prenatal: -0.08
(-0.20, 0.04)e
Maternal postnatal: 0.03
(-0.08, 0.13)e
Child: -0.06 (-0.19, 0.06)e
Males
Maternal prenatal: -0.06
(-0.23, 0.11 )e
Maternal postnatal: -0.01
(-0.14, 0.13)e
Child: -0.08 (-0.25, 0.10)e
Females
Maternal prenatal: -0.04
(-0.19, 0.11 )e
Maternal postnatal: 0.07
(-0.08, 0.21 )e
Child: -0.05 (-0.21, 0.11)e
Total Adaptive Skills
Maternal prenatal: -0.06
(-0.19, 0.07)e
Maternal postnatal: 0.08
(-0.03, 0.19)e
Child: 0.08 (-0.06, 0.22)e
Males
Maternal prenatal: -0.01
(-0.19, 0.18)e
Maternal postnatal: 0.08
(-0.07, 0.24)e
3-447
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StillyD^kjn* Study Population Exposure Assessment Outcome Confounders Effect Estimates and 95% CIs
Child: 0.10 (-0.13, 0.32)e
Females
Maternal prenatal: -0.19
(-0.34, —0.04)e
Maternal postnatal: 0.07
(-0.08, 0.23)e
Child: 0.26 (0.07, 0.45)e
tZhou etal. (2017)
Shanghai
China
2010-2012
Followed through 24-
36 mo
Cohort
Shanghai Stress
Birth Cohort study
n: 139
Mother-infant pairs
in prenatal clinics
of maternity
hospitals during
mid-to-late
pregnancy
Blood
Maternal whole blood
Age at measurement: wk 28-36
of gestation
GM (95% CI): 3.30 (3.05, 3.57)
pg/dL
Adaptive and social
behavior domain DQs
from GDS
Age at outcome: 24-
mo
¦36
Maternal age at
enrollment, SES,
maternal education,
gestational week, child
sex, birth weight and
age
Beta per log—10 transformed
BLL
Adaptive:
Overall: 3.60 (-3.64, 10.83)b
Low stress: 7.57 (-0.12,
15.27)b
High stress: -17.93 (-35.83,
—0.03)b
Social:
Overall: -6.45 (-15.55, 2.65)b
Low stress: -0.07 (-9.57,
9.44)b
High stress: -41.00 (-63.11,
—18.89)b
3-448
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tRuebner et al.
(2019)
46 centers
U.S.
Study Years: NR
Cross-sectional
CKiD Cohort study
n: 412
Children (age 1-
yr) with mild to
moderate CKD
¦16
Blood
Child venous blood; ICP-MS. The
BLL measurement closest to the
time of neurocognitive testing
was used for analysis
(concurrent).
Age at measurement:
NR; 2, 4, or 6 yr after study entry
Median: 1.2 pg/dL
75th: 1.8 Mg/dL
Max: 5.1 |jg/dL
Adaptive skills,
composite index on the
BASC-2 (see also 3.5.1
and 3.5.2)
The last available test
results were used to
evaluate long-term
effects. Mean time
between BLL and
neurocognitive testing
was 2.3 yr.
Age at outcome:
3, 5, or 7 yr after study
entry
Child age, sex, race,
poverty, and maternal
education
Adjusted BASC-2 results were
not reported because they
were not statistically significant.
tViqeh et al. (2014)
Tehran
Iran
October 2006 -
March 2011
Followed through 36
mo
Cohort
Birth cohort
n: 174
Mother-infant pairs
recruited in first
trimester (8-12
wk).
Blood
Maternal blood, cord blood;
MS
ICP-
Age at measurement:
3 trimesters during pregnancy
and delivery
Mean: 1st trimester: 4.15 |jg/dL,
2nd trimester: 3.44, 3rd trimester:
3.78, umbilical cord: 2.86
Max: 1st trimester: 20.5 |jg/dL,
2nd trimester: 7.5, 3rd trimester:
8.0, umbilical cord: 6.9
Mental development
assessed using the
ECDI by Harold Ireton
(language
comprehension,
expressive language,
gross motor, self-help,
social interaction).
Cutoff point scores for
development delay was
score <20% of that
expected for children's
age.
Age at outcome:
36 mo
Maternal educational,
BMI, family income,
gestational age, birth
weight, birth order (first
born)
OR
Total ECDI:
1.74 (1.18, 2.5)
3-449
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tKimetal. (2018b)
4 cities: Seoul,
Anyang, Ansan and
Jeju
Korea
Pregnancy (2011-
2012) through 24 mo
of age
Cohort
CHECK cohort
n: 140
birth cohort-
pregnant women
recruited from 4
cities in Korea
before delivery,
Blood
Prenatal maternal blood collected
during hospital visit: 2.7 |jg/dL
Cord blood: 1.2 pg/dL
Adaptive behaviors
assessed using SMS
Association was
examined using multiple
linear regression
analysis.
Age at outcome:
13-24 mo
BPA, and phthalates, Associations of blood Pb
maternal age
(continuous), birth
delivery mode
(categorical), monthly
household income
(categorical), child's
sex, and BDI
(continuous) of the
mother, gestational
age (continuous),
primiparous
(categorical), and
pre-pregnancy BMI
(categorical)
concentrations and SQ were
assessed but not reported
because they lacked statistical
significance.
AAS = atomic absorption spectrometry; ADHD = attention deficit/hyperactivity disorder; ADOS = Autism Diagnostic Observation Schedule; ASD = autism spectrum disorder; ASQ =
Ages and Stages Questionnaire Inventory; ASSQ = Autism Spectrum Screening Questionnaire; BASC = Behavior Assessment System for Children; BDI = Beck Depression
Inventory; BLL = blood lead level; BMI = body mass index; BPA = bisphenol A; BRS = behavioral rating scale; BSID = Bayley Scales of Infant and Toddler Development; CARS =
Childhood Autism Rating Scale; CDIIT = Comprehensive Developmental Inventory for Infants and Toddlers; CHECK = Children's Health and Environmental Chemicals in Korea;
CHEER = Children's Health and Environmental Research; CKiD = Chronic Kidney Disease in Children; DQ = development quotient; DSM = Diagnostic and Statistical Manual of
Mental Disorders; GM = geometric mean; ECDI = Early Child Development Inventory; ELEMENT = Early Life Exposure in Mexico to Environmental Toxicants; GDS = Gesell
Developmental Schedules; GFAAS = graphite furnace atomic absorption spectrometry; HOME = Home Observation Measurement of the Environment; ICP-MS = inductively coupled
plasma mass spectrometry; ISAT = Illinois Standard Achievement Test; MAT = Metropolitan Achievement Test; MEAP = Michigan Educational Assessment Program; MDAT =
Malawi Development Assessment Tool; Mn = manganese; mo = month(s); NHANES = National Health and Nutrition Examination Survey; NHBCS = New Hampshire Birth Cohort
Study; NHNPR = Norwegian Patient Registry; NR = not reported; OR = odds ratio; Pb = lead; SD = standard of deviation; SES = socioeconomic status; SGA = small for gestational
age; SMS = Social Maturity Scale; SRS = Social Responsiveness Scale; SQ = social quotient; TBPS = Taiwan Birth Panel Study; wk = week(s); yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResults are unstandardized because the Pb level distribution data was not available.
The CI was calculated from a p-value and the true CI may be wider or narrower than calculated.
dResults are unstandardized because the log base used for exposure transformation was unspecified in the study.
eResults are unstandardized because the biomarker used for Pb exposure measurement is not blood, tooth, or bone.
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-450
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Table 3-14E Epidemiologic studies of exposure to Pb and cognitive function in adults
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tPower et al. (2014)
Boston, MA, U.S
1993-2008
Cohort
Participants selected from
cohort study (Nurse's Health
Study) and part of case-
control ancillary study
n: 584
Bone, Blood
Bone Pb: K-XRF at the
midtibial shaft and the
patella, blood Pb
concentrations; GFAAS with
Zeeman background
correction in year 1993-
2004
Age at measurement:
registered nurses aged 45-
74 yr
Tibia Pb cone: 10.5 ± 9.7
|jg/g, Patella Pb cone: 12.6 ±
11.7 pg/g.
Blood Pb cone: 2.9 ± 1.9
pg/dL
Cognitive decline
Cognitive decline
assessed using a
telephone battery
of cognitive tests
during 2-4 waves
over the period of
follow-up, 1995-
2008. All 9
cognitive scores
were Z-transformed
with high score
representing better
performance.
Alcohol
consumption,
smoking status,
education,
husband's
education,
menopausal
status/hormone
therapy use,
physical activity,
ibuprofen use,
aspirin use, vitamin
E supplementation,
the % of residential
census tract of
white race/ ethnicity,
and median income
of residential
census track.
Beta (95% Cl)a
Tibia
Verbal Memory
-0.002 (-0.006, 0.003)
Overall Cognition
-0.002 (-0.005, 0)
Patella
Verbal Memory
-0.001 (-0.005, 0.002)
Overall Cognition
-0.001 (-0.004, 0.001)
Blood
Verbal Memory
0.003 (-0.021, 0.027)
Overall Cognition
-0.007 (-0.023, 0.009)
3-451
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tFarooaui et al. (2017)
Boston, MA, U.S.
1993-2007
Cohort
Participants selected from
cohort study (Veterans Affairs
NAS
n: 741 subjects in MMSE and
715 in Global cognition
Bone
Patella (trabecular bone)
and tibia (cortical bone)
bone Pb was measured
using K-XRF spectroscopy in
1993
Age at measurement:
healthy men aged 51-98 yr
Patella Pb cone: 30.6 ±
19.44 |jg/g, and tibia Pb
cone: 21.6 ± 13.33 |jg/g
Changes in
cognition
Cognition was
assessed using the
MMSE, NES2,
CERAD and WAIS-
R during 3-5 visits
over the period of
15 yr of follow-up.
Age at first cognitive
test, past education
level, baseline
smoking status and
alcohol intake.
Beta (95% Clf Pb and MMSE
over time
Tibia
IQR change in Pb -0.051
(-0.137, 0.035)
IQR change in Pb*time -0.007
(-0.018, 0.004)
Patella
IQR change in Pb -0.061
(-0.12, -0.002)
IQR change in Pb*time -0.008
(-0.015, 0)
HR (95% Cl)b
Patella 1.095 (0.993, 1.207)
Tibia 1.033 (0.875, 1.22)
Beta (95% Cl)b Pb and Global
Cognition overtime
Patella -0.119 (-0.247, 0.009)
Tibia -0.137 (-0.318, 0.043)
3-452
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tWeuve etal. (2013)
Boston, MA, United
States
2003-2007
Cross-sectional
PD cases confirmed by
movement disorder specialists
using the U.K. Brain Bank
criteria
n: 151 subjects (101 cases
and 50 controls)
Bone
Bone Pb measured using K-
XRF spectrometric estimates
of Pb concentrations in Tibia
and Patella bones.
Age at measurement: cases
and controls (spouses, in-
laws, or friends of the cases)
aged 54-81 yr
Patella Pb cone by age at
cognitive interview
categories:
54-64.9 yr
65-69.9 yr
70-74.9 yr
75-80.9 yr
5.9 ± 10.3 |jg/g
9.2 ± 7.8 |jg/g
7.7 ± 10.5 |jg/g
15.2 ± 10.2 |jg/g
Tibia Pb cone by age at
cognitive interview
categories:
54-64.9 yr
65-69.9 yr
70-74.9 yr
75-80.9 yr
4.4 ± 11.1 |jg/g
8.8 ± 10.5 |jg/g
6.8 ± 8.8 |jg/g
9.2 ±11.5 |jg/g
Cognition function Age at cognitive
assessment, sex,
Cognitive function race, education,
assessed using a smoking history
telephone cognitive
assessment battery
of 9 tests based on
a validated
telephone battery
for assessing age-
related cognitive
decline. Added test
of cognitive
domains that
typically decline in
PD. All 9 cognitive
scores were z-
transformed with
high score
representing better
performance.
Adjusted difference (95% Cl)b
Patella
Telephone interview for
cognitive assessment (TICS)
-0.08 (-0.32 to 0.15)
Delayed 10-word recall
0.05 (-0.18 to 0.28)
Delayed 10-word recognition
0.01 (-0.22 to 0.24)
Animal naming
-0.11 (-0.32 to 0.10)
"F" naming
-0.07 (-0.30 to 0.17)
Digit span forward
-0.02 (-0.27 to 0.22)
Digit span backward
0.05 (-0.17 to 0.27)
Oral trails B minus A
0.03 (-0.23 to 0.28)
Global score
-0.01 (-0.14 to 0.13)
Tibia
Telephone interview for
cognitive status (TICS)
-0.20 (-0.40 to -0.00)
Delayed 10-word recall
-0.04 (-0.23 to 0.16)
Delayed 10-word recognition
-0.01 (-0.21 to 0.20)
Animal naming
-0.11 (-0.29 to 0.07)
"F" naming
-0.19 (-0.39 to 0.01)
Digit span forward
3-453
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Referenc^and Study Study Population Exposure Assessment Outcome Confounders Effect Estimates and 95% CIs
-0.23 (-0.43 to -0.03)
Digit span backward
-0.19 (-0.37 to -0.00)
Oral trails B minus A
-0.06 (-0.29 to 0.17)
Global score
-0.13 (-0.25 to -0.01)
tSkerfvina et al. (2015) n: 927
Landskrona and
Trelleborg, Southern
Sweden
1978-2007 followed for
4-12 yr
Cohort
Blood
Between 1978 and 1994, B-
Pb levels were determined
using flame or
electrothermal atomization
atomic absorption
spectrometry; between 1995
and 2007, B-Pb levels were
determined using inductively
coupled plasma mass
spectrometry
Age at measurement: 7-12
yr
IQ assessed for
military
conscription
IQ (measured
logical, verbal,
spatial abilities, and
technical
understanding)
assessed as a part
of military
conscription
examinations.
Age at outcome:
18-19 yr
Age at blood
sampling, sex,
parents' education,
family economy,
and country of birth
of child and parents
Beta (SE)a
IQ
All subjects
-0.127 (-0.209, -0.045)
Blood Pb <50 |jg/L
-0.204 (-0.392, -0.016)
Mean: 34 |jg/L
3-454
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tReuben et al. (2017) Dunedin Multidisciplinary
Health and Development
Dunedin, New Zealand Study
1972/73-2012
Cohort
n: 565
Blood
Graphite fumance atomic
absorption
spectrophotometry
Age at measurement: 11 yr
Mean (SD): 10.99 ±4.63
pg/dL
Full -scale IQ
(other domains
such as verbal
comprehension,
perpetual
reasoning, working
memory,
processing speed)
Cognitive function
assessed using
Wechsler Adult
Intelligence Scale -
IV (WAIS-IV) at the
age of 38 yr.
Childhood IQ scores Change in IQa (95% CI)
(age 7 and 9 yr),
their mothers' IQ
score, and their
socioeconomic
background
Adjusted by sex
-0.394 (-0.669, -0.119)
Fully adjusted
-0.322 (-0.496, -0.148)
Change in perceptual
reasoning3 (95% C)
-0.414 (-0.627, -0.201)
Change in working memory3
(95% CI)
-0.252 (-0.476, -0.028)
Change in socioeconomic
status3 (95% CI)
-0.358 (-0.635, -0.081)
tReuben et al. (2020) Dunedin Multidisciplinary
Health and Development
Dunedin, New Zealand Study
1972/73-2019
Cohort
n: 564
Blood
Furnace atomic absorption
spectrophotometry
Age at measurement: 11 yr
Mean (SD): 10.99 ±4.63
pg/dL
Full-scale IQ and
self-reported
information)
Cognitive
performance
assessed
objectively using
Wechsler Adult
Intelligence Scale -
IV (WAIS-IV) and
subjectively via
informant and self-
reports at the age
of 45 yr.
Childhood IQ scores
(age 7 and 9 yr),
their mothers' IQ
score, and their
socioeconomic
background
Change in IQ3(95% CI)
-0.414 (95% CI: -0.679,
-0.149)
Residualized Change in IQ3
(95% CI)
-0.394 (-0.583, -0.205)
3-455
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tKhalil etal. (2014)
Multicity (6 clinical
sites), U.S.
May 2007 to Nov 2008
Cross-sectional
Population-based cohort study Blood
(MrOS)
n: 445 Venous blood samples
tested for Pb levels using
AAS
Age at measurement: non-
Hispanic Caucasian men
aged >65 yr
Mean (SD)
2.25 ± 1.20 |jg/dL
Cognitive function
assessed using 3
MS
Age at outcome:
>65 yr
Age, education,
smoking, alcohol
consumption and
BMI
Beta (95% Cl)b
Cognitive Function in Adults
3MS
-0.01 (-1.10,1.07)
Cognitive Function in Adults
Trail Making B
2.72 (-7.65, 13.09)
tSouza-Talarico et al.
(2017)
Sao Paulo City,
Brazil
Cross-sectional
n: 125 (104 women and 21
men)
Blood
Venous blood samples
tested for heavy metals (Cd
and Pb) levels using ICP-MS
Age at measurement:
Healthy older adults
between 50 and 82 yr (M =
65.9)
Mean (SD)
2.1 ± 0.970 |jg/dL
MMSE and
Informant
Questionnaire on
Cognitive Decline
used to rule out
cognitive and
functional
impairments.
Age at outcome: 50
and 82 yr (M =
65.9)
Age, sex, income,
education,
hemoglobin,
hematocrit
Betab
WMC Pb 0.106 (AR: 0.057)
BCd x BPb interaction-term and
WMC-0.378 (p< 0.001)
Table 3 adjusted EE for Pb and
WMC, with and without
controlling for Oxygen Radical
Absorbance Capacity total:
standardized
3-456
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tvan Wiinaaarden et al.
(2011)
Nationwide, U.S.
1999-2008
Cross-sectional
NHANES 1999-2008 (for self-
reported confusion and
memory problems) and
NHANES 1999-2002 (for
DSST)
n: 9526 participants (7277
from NHANES 1999-2008
and 2299 participants from
NHANES 1999-2002)
Blood
Venous blood samples
tested for Pb concentration
using AAS with Zeeman
background correction.
Age at measurement: >60 yr
Blood Pb cone: 2.46 |jg/dL
(range 0.18-54.00 pg/dL)
Cognitive function
Cognitive function
assessed by self-
reported responses
on limitation in
cognitive
functioning, and
DSST (a subset of
the WAIS-III) for
subset of
participants.
Age, sex, ethnicity,
education level,
PIR, self-reported
general health
status
OR (95% Cl)b
1.01 (0.65, 1.56)
Age at outcome:
>60 yr
tPrzvbvlaet al. (2017) NHANES cycles 1999-2000 Blood
Nationwide, U.S.
1999-2002
Cross-sectional
and 2001 -2002;
n: 498
Cognitive function Race/ethnicity, age, Betas per natural log increase in
Blood samples tested for
chemicals (Pb, Cd and
PCBs) concentrations; Pb
and Cd measured using
ICP-MS.
Age at measurement: 60-84
yr
Mean 2.17 pg/dL (95% CI:
2.07, 2.27)
Cognitive function
assessed using the
DSC Module of the
WAIS-III.
Age at outcome:
60-84 yr
education level,
PIR, sex and
smoking status
BLL
Cognitive Functioning
All Participants: -0.10 (-0.20,
-0.01)
Females:
-0.12 (-0.26, 0.01)
Males:
-0.09 (-0.24, 0.06)
Age 60-69:
-0.13 (-0.28, 0.01)
Age 70-74:
-0.08 (-0.2, 0.04)
3-457
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tSasaki and Carpenter NHANES cycles 2011-12 and Blood and Urine
(2022)
Nationwide, U.S.
2011-2014
Cross-sectional
2013-14 and tested for
different sets of chemicals for
different subgroups
n: 3042
Venous blood samples and
urine samples tested for
seven metals and metalloids
(including Pb) using ICP-MS
Age at measurement: 60-80
yr
Blood mean Pb: 19.0 |jg/L
Urine mean Pb: 0.72 |jg/dL
Cognitive function Age, sex, ethnicity,
education level,
Immediate and depression,
delayed memory diabetes, alcohol
assessed using the consumption, and
CERAD, and smoking
working memory
assessed using the
DSST.
Age at outcome:
60-80 yr
Beta (95% Cl)b
Blood
CERAD Immediate recall:
-0.58 (-0.91, -0.24)
CERAD Delayed recall:
-0.19 (-0.35, -0.02)
Digit symbol substitution:
-1.08 (-2.12, -0.05)
CERAD immediate recall as a
function of age
60s Years Old Group:
-0.37 (-0.87, 0.13)
>70 Years Old Group:
-0.85 (-1.44, -0.27)
Urine
CERAD Immediate recall:
-0.26 (-0.58, 0.06)
CERAD Delayed recall:
-0.03 (-0.19, 0.13)
Digit symbol substitution:
-1.03 (-2.01, -0.06)
tXiao etal. (2021)
Guangxi, southern
China
Aug 2016-July 2018
Cross-sectional
n: 2879
Blood
Venous blood samples
tested for 22 metals
(including Pb) using ICP-MS.
Age at measurement: >60 yr
Blood Pb: Median: 51.5 |jg/L
Cognitive function Age, gender,
Cognitive function
assessed using the
MMSE.
Age at outcome:
>60 yr
education
attainment, annual
income, BMI,
smoking, alcohol
drinking, insomnia,
and physical activity
Beta (95% Cl)a
Cognitive function
Single-pollutant model
-0.018 (-0.06, 0.023)
Multi-pollutant model
-0.019 (-0.063, 0.025)
3-458
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tMeramat et al. (2017) Neuroprotective Model for
Healthy Longevity among
Malaysia
May 2013 to January
2014
Cross-sectional
Malaysia Older adult
n: 317
Nail
Toenails (clipped from all
toes) assessed for trace
elements (Al, Ca, Cd, Co,
Fe, Pb, Zn, Se, Cu and Cr)
using ICP-MS.
Age at measurement: >60 yr
Pb cone: Cognitive impaired
group (n = 197): 0.55 ± 0.03
|jg/g; and Normal cognitive
group (n = 120): 0.35 ±
0.013 |jg/g
Cognitive
impairment
assessed using
Montreal Cognitive
Assessment - a
Malay version
Age at outcome:
>60 yr
Age, sex, years of
education and
smoking habits
OR (95% Cl)a
Cognitive impairment 2.471
(1.535-3.980)
3-459
-------
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
tYu et al. (2021)
Nationwide, U.S.
Jan 2015-Sep 2017
Cohort
SPHERL longitudinal study Blood
n: 260 (260: DSST cohort and
168: SCWT cohort) with
baseline and annual follow-up
blood Pb measurements and
neurocognitive function
assessments.
Venous blood samples
tested for Pb concentration
using ICP-MS.
Age at measurement: mean
age 29.4 yr
Blood Pb cone: DSST
cohort: Geo mean: 3.97 (5-
95th percentage interval (PI)
0.90-14.3) |jg/dL at
baseline, 13.4 (PI 3.70-30.3)
|jg/dL and 12.8 (PI 2.80-
29.2) |jg/dL at the first and
second follow-up visits,
respectively.
Cognitive function
Cognitive function
changes assessed
using the DSST
and ST at baseline
and annual follow-
up visits.
Age at outcome:
mean age 29.4 yr
Age, sex, ethnicity,
change in age,
baseline BMI,
changes in body
weight, education,
baseline blood Pb,
baseline
neurocognitive
function test,
baseline values and
changes in smoking
status, total/HDL
ratio, cholesterol
and alcohol
consumption
OR (95% Cl)a
DSST
1.012 (0.997, 1.028)
3MS = Modified Mini Mental State Examination; BrainAGE = Brain Age Gap Estimation; CERAD = Consortium to Establish a Registry for Alzheimer's Disease; DSC = Digital Symbol
Coding; DSST = Digit Symbol Substitution Test; EE = effect estimate(s); K-XRF = K-shell X-ray fluorescence; MMSE = mini mental status exam; MrOS = Osteoporotic Fractures in Men
Study; NAS = Normative Aging Study; NES2 = Neurobehavioral Evaluation System 2; PD = Parkinson's disease; WAIS-III = Wechsler Adult Intelligence Scale, Third Edition; WAIS-R =
Wechsler Adult Intelligence Scale-Revised; WMC = working memory capacity; SPHERL = Study for Promotion of Health in Recycling Lead; SCWT = Stroop Color-Word Test.
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a change
in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResult not standardized because data pertaining to the BLL distribution and/or base for the log-transformation were not reported.
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-460
-------
Table 3-15E Epidemiologic studies of Pb exposure and psychopathological effects in adults
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
Raian et al. (2007)
Boston, MA, U.S.
1991-2002
Cohort
Veterans Affairs
NAS
n: 1,075
Closed cohort of
male volunteers with
no chronic medical
conditions at entry.
97% white
Bone
Bone Pb measured in the mid-
tibia shaft and patella using K-
XRF
Age at measurement: 21-80 yr
Mean: -67.5 yr old
Mean (SD):
Tibia: 22.1 (13.8) pg/g
Patella: 31.4 (19.6) pg/g
Depression and anxiety
Depressive and anxiety
symptoms were measured using
the BRIEF Symptom Inventory
(depression and anxiety were
determined to be present for
participants that scored 1 SD
above the mean for a normal
population). Participant followed
up was 3 yr.
Age, alcohol
consumption,
education, time
between
assessments, and
cumulative smoking
Anxiety OR (95% Cl)a
Tibia: 1.13 (0.99, 1.29)
Patella: 1.09 (0.99, 1.19)
Depression OR (95% Cl)a
Tibia: 1.11 (0.98, 1.38)
Patella: 1.05 (0.96, 1.16)
Bouchard et al.
(2009)
U.S.
1999-2004
Cross-sectional
NHANES
n: 1,987
Blood
Blood Pb measured in venous
whole blood samples using
ICP-MS
Age at measurement:
20-39 yr old
Geo. mean: 1.24 pg/dL
20th %ile: 0.7 pg/dL
40th %ile: 1.0 pg/dL
60th %ile: 1.4 pg/dL
80th %ile: 2.1 pg/dL
Depression
WHO CIDI was administered.
Major depressive disorder
diagnosed according to DSM-IV
criteria.
Age at outcome: 20-39 yr
Age, sex,
race/ethnicity,
education, and PIR
Major Depressive Disorder
OR (95% Cl)a
Q1
Q2
Q3
Q4
Q5
2.72)
Ref.
1.39 (0.71,
1.28 (0.69, 2.38)
1.41 (0.76, 2.6)
2.32 (1.13, 4.75)
3-461
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tPeters et al. (2011)
Boston, MA
U.S
1991-1997(Bone Pb
measurements);
1993-2003
(Psychological
measurements)
Cohort
Veterans Affairs
NAS
n: 412
Closed cohort of
male volunteers with
no chronic medical
conditions at entry.
97% white
Bone
Bone Pb measured in the mid-
tibia shaft using K-XRF
Age at measurement:
Mean: -65.3 yr old
Mean: 20.6 |jg/g
Pessimism and Depression
A subscale of the Life
Orientation Test was used to
assess pessimistic attitudes.
Depressive symptoms were
measured using the BRIEF
Symptom Inventory (depression
was determined to be present
for participants that scored 1 SD
above the mean for a normal
population).
Age at outcome: Mean -68.3 yr
Age, health
behaviors,
childhood SES,
adult SES
Difference in Pessimism
Level on the Life Orientation
Test (95% CI)
0.21 (0.00, 0.43)
tReuben et al.
(2019)
Dunedin, New
Zealand
Enrollment: 1972-73;
Follow-up through
2012
Cohort
Dunedin
Multidisciplinary
Health and
Development Study
Cohort of children 3
yr old at enrollment
followed through 32
yr of age. Study
population was
nationally
representative
(majority white) and
had high rates of
participation and
follow-up.
Blood
Blood Pb measured in venous
blood samples using GFAAS
Age at measurement: 11 yr
Mean: 11.08 pg/dL
(94% above 5 |jg/dL)
General Psychopathology,
Externalizing Symptoms,
Internalizing Symptoms, and
Thought Disorder Symptoms in
Adults
Psychopathology symptoms
were assessed using the
Diagnostic Interview Schedule.
Factor loadings from each of 11
disorders were used to create
hierarchical measures for
psychopathology and each of its
constituent psychiatric spectra
Sex, childhood
maternal IQ, and
family history of
mental illness.
Change in symptom scores
(95%CI) (standardized to a
mean [SD] of 100 [15])a
General Psychopathology
0.27 (0.02, 0.51)
Externalizing Symptoms
0.15 (-0.10, 0.40)
Internalizing Symptoms
0.28 (0.04, 0.53)
Thought Disorder
0.26 (0.01, 0.51)
Age at outcome:
and 38 yr
18, 21, 26, 32,
3-462
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tMcFarlane et al.
(2013)
Port Pirie, Australia
1979-1982
(enrollment); 2008-
2009 (follow-up)
Cohort
Port Pirie cohort
Study
n: 210
Mother-singleton
infant pairs enrolled
in Pb-smelting town
from 1979-1982.
Assessed
periodically from
birth to 7 yr, again
from 11 to 13 yr, and
for this study, at 25
to 29 yr
Blood
Blood Pb measured in capillary
blood samples using GFAAS
Age at measurement:
6, 15, and 24 mo; 3-7 yr
Mean: 17.2 |jg/dL
(birth to 7-yr average)
Drug and alcohol abuse,
DSM-IV Disorders (Alcohol
abuse, drug abuse, social
phobia, specific phobia, PTSD,
alcohol dependence, panic
attack, major depressive
disorder) and adult self-report
DMV-IV oriented subscale
(anxiety, somatic problems,
depressive problems,
hyperactivity, inattention,
antisocial personality problems,
avoidant personality problems)
Age at outcome:
25 to 29 yr
HOME, maternal
education, paternal
occupation,
mothers' age at
birth, breastfeeding,
and single parent
family status
OR (95% Cl)a
Social Phobia
Women: 1.05 (0.93, 1.188)
Men: 0.96 (0.80, 1.15)
Specific Phobia
Women: 1.13 (0.99, 1.29)
Men: 1.02 (0.71, 1.47)
Major Depressive Disorder
Women: 0.89 (0.77, 1.03)
Men: 0.89 (0.68, 1.16)
3-463
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tLi etal. (2017)
n: 1,701
Shanghai (inner and Stratified cluster
outer districts)
China
2010
Cross-sectional
sampling of
pregnant women
(gestational wk 28-
36)
Blood
Maternal stress
Blood Pb measured in venous Life Event Stress Scale for
blood samples using GFAAS
Geo mean: 3.97 |jg/dL
Max: 14.84 pg/dL
Pregnant Women, Symptom
Checklist-9-Revised (GSI
[measure of psychological
distress], anxiety and
depression scores)
Age at outcome:
13-42 yr old (wk 28-36)
Maternal age at
enrollment,
ethnicity, maternal
education, and
family monthly
income, years
residing in Shanghai Maternai stress
Change in maternal stress,
anxiety, and depression
scores per 10-fold increase in
BLLs (results from piecewise
linear models)3
<2.57 |jg/dL: 0.22 (0.05, 0.4)
>2.57 |jg/dL: -0.07 (-0.16,
0.01)
Depression
<2.57 |jg/dL: 0.34 (0.12,
0.56)
>2.57 |jg/dL: -0.09 (-0.19,
0.02)
Anxiety
<2.57 |jg/dL: 0.25 (0.04,
0.46)
>2.57 |jg/dL: -0.08 (-0.18,
0.02)
tlshitsuka et al.
Japan Environment
Blood
Maternal Depression
Age, parity, marital
OR (95% Cl)b
(2020)
and Children's Study
status, education,
n: 17,267
Blood Pb measured in whole
K6. Depression measured as
employment status,
K6 >13
Japan
blood samples using ICP-MS
scores >5 or 13 (two cutoff
household income,
1.00 (0.76, 1.32)
Pregnant women
Age at measurement:
points for sensitivity).
and smoking and
2011-2014
recruited out of 15
31 yr (mean)
alcohol status
regional centers
Age at outcome:
K6 >5
Cross-sectional
across Japan
Geo. mean: 0.58 |jg/dL
Max: 6.75 |jg/dL
Mean age: 31 yr
0.98 (0.88, 1.09)
3-464
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tBerk et al. (2014)
NHANES
n: 15,140
Blood
Depression
Age, sex, poverty,
family income,
Depression OR (95% CI)
U.S.
Blood Pb measured in venous
Depression measured as >9 on
ethnicity, and
Q4 vs. Q1*:
General population,
whole blood samples using
the nine-item depression
country of birth
0.98 (0.78, 1.25)
2005-2010
>18 yr old
ICP-MS
module of the Patient Health
Age at measurement:
Questionnaire
>18 yr old
Mean: NR
*Quartile levels NR
Cross-sectional
Age at outcome:
>18 yr old
tNauven et al. (2022) KNHANES
n: 16,371
South Korea
2009-2013 and
2016-2017
Cross-Sectional
General population;
mean age: 42.6 yr
old (SD: 18.12)
Blood
Blood Pb was measured in
venous whole blood using
GFAAS
Age at measurement (mean):
42.6 yr old (SD: 18.12)
Geo. Mean:
1.84 |jg/dL (w/o depression)
1.85 |jg/dL (w/ depression)
Depression
Self-reported physician's
diagnosis or treatment for
depression
Age at outcome (mean): 42.6 yr
old (SD: 18.12)
Sex, urbanicity,
household income,
physical activity,
occupation, BMI,
alcohol
consumption,
education level, and
smoking status
OR (95% Clf
1.02 (0.90, 1.16)
3-465
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tEumetal. (2012)
Boston, MA
United States
Subsample 1: Bone
Pb Measure 1993-
1995;
Subsample 2: Bone
Pb Measure 2001-
2004.
Psychological
Questionnaires:
1988, 19992,1996,
2000, 2004
Cohort
Nurses' Health
Study
n: 617
Women from two
subsample studies
of the NHS cohort
Bone
Midtibial shaft and patella bone
Pb measured using K-XRF
Age at measurement:
Mean: 60.9 yr
Mean:
Tibia: 10.3 |jg/g;
Patella:12.5 |jg/g
Tibia Tertiles:
T1: <7.0 |jg/g
T2: 7.0-11.5 pg/g
T3: >11.5 pg/g
Patella Tertiles:
T1
T2
T3
<8.5 pg/g
8.5-14.5 pg/g
>14.5 pg/g
Phobic anxiety and depressive
symptoms
Depression symptoms
measured using MHI-5; Anxiety
symptoms measured using
phobic anxiety scale of the
Crown-Crisp Experiential Index
(CCEI)
Age at outcome:
Mean:
MHI-5: 59.4 yr
CCEI: 59.2 yr
Substudy group,
age at bone Pb
measure, age at
MHI-5 or CCEI
measurement,
education,
husband's
education, alcohol
consumption, pack-
years of smoking,
and employment
status at MHI-5 or
CCEI assessment
OR (95% Cl)b (Tertile 3 vs.
Tertile 1)
CCEI >4
All women
Tibia: 1.10 (0.73, 1.64)
Patella: 0.75 (0.49, 1.15)
Women on HRT
Tibia: 2.79 (1.02, 7.59)
Patella: 0.23 (0.07, 0.69)
MHI-5 Point Difference
(lower scores indicate worse
symptoms)
All women (T3 vs. T1)
Tibia: -1.06 (-3.05, 0.94)
Patella: -7.78 (-11.73,
-3.83)
Women on HRT (T3 vs. T1)
Tibia: 0.61 (-1.55, 2.78)
Patella: 0.51 (-3.91, 4.94)
3-466
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tFan et al. (2020) Cohort Study of
Luan city, Anhui
province, China
2016
Cross-sectional
Elderly Health and
Environmental
Controllable Factors
n: 994
Older adults (>60 yr
old) selected using
cluster sampling
from two
communities in
Luan, China
Blood
Blood Pb measured in venous
whole blood samples using
ICP-MS
Age at measurement:
>60 yr old
Quartiles
Q1
Q2
Q3
Q4
<2.03 |jg/dL
2.03-2.68 |jg/dL
2.68-3.06 |jg/dL
>3.06 |jg/dL
Depressive symptoms
Chinese revision of the geriatric
depression scale
Age at outcome:
>60 yr old
Age, gender, region, OR (95% Cl)b
marital status,
monthly income,
education level,
alcohol intake,
smoking, and BMI
Depression
Q1
Q2
Q3
Q4
Ref.
1.28 (0.79,
1.36 (0.84,
2.03 (1.23,
2.08)
2.22)
3.35)
tMaetal. (2019)
Hebei Province
China
2018-2019
Case-control
n: 190 (95 cases,
controls)
95 Blood
First-episode drug-
naive patients ages
18 to 60 yr old were
recruited from a
psychiatric hospital.
Age and sex-
matched controls
without known
psychiatric problems
recruited from an
affiliated hospital
Serum Pb measured in venous
blood samples using ICP-MS
Age at measurement:
18-60 yr old
Median: 0.61 ng/mL (serum)
75th: 0.79 ng/mL (serum)
Schizophrenia
Physician-diagnosed
schizophrenia using ICD-10
criteria
Age at outcome:
18-60 yr old
Marital status
(others not
specified).
Population matched
on age and sex
OR (95% Cl)b per 1 ng/mL
increase
3.15 (1.24, 7.99)
AAS = atomic absorption spectrometry; BLL = blood lead level; BMI = body mass index; CCEI = Crown-Crisp Experiential Index; CI = confidence interval; CIDI = Composite
International Diagnostic Interview; GFAAS = graphite furnace atomic absorption spectrometry; K6 = Kessler Psychological Distress Scale; K-XRF = K-shell X-ray fluorescence; MHI-5
= Mental Health Index 5-item; NAS = Normative Aging Study; NR = not reported; Pb = lead; PIR = poverty-income ratio; PTSD = post-traumatic stress disorder; Q = quartile; SD =
standard deviation; SES = socioeconomic status; WHO = World Health Organization; wk = week(s); yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResult not standardized because data pertaining to the BLL distribution and/or base for the log-transformation were not reported.
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-467
-------
Table 3-16E Epidemiologic studies of Pb exposure and sensory organ function in adults
RefereDCesfgnnd Study Population Exposure Assessment Outcome Confounders Effect f|o^eS a"d
Park etal. (2010)
Eastern Massachusetts
U.S.
Enrollment and outcome
assessment: 1962-1996;
bone Pb measurements:
1991-1996
Cohort
NAS
n: 448
Bone
Bone Pb levels measured
in the midtibial shaft and
patella with a K-XRF
instrument.
Age at measurement:
Mean (SD) at bone Pb
measurement = 64.9 (7.3)
yr; mean (SD) at first
audiometric test = 42.5
(8.4) yr
Mean (SD) in tibia = 22.5
(14.2) |jg/g; mean (SD) in
patella = 32.5 (20.4) pg/g
Sensory Organ Function
Pure-tone averages
assessed by audiologists
with the modified Hughson-
Westlake procedure. Air
conduction hearing
thresholds measured for
each ear by audiologists
using either a Beltone 15C
or a Grason-Stadler 1701
audiometer.
Cross-sectional
analyses and logistic
regression analyses
adjusted for age,
race, education, BMI,
pack-years of
cigarettes, diabetes,
hypertension,
occupational noise,
and noise notch.
Hearing loss OR (95%
Cl)b
Tibia 1.19 (0.92, 1.53)
Patella 1.48 (1.14, 1.91)
EE in Hearing thresholds
(dB HL) with one
interquartile range
Increment in bone lead
measure
Tibia PTA 0.83 (-0.18,
1.83)
Patella PTA 1.58 0.62,
2.55
tShiue (2013)
U.S.
2003-2004
Cross-sectional
NHANES
n: 712 (vision); 732
(hearing); 669
(balance)
NHANES age 50 and
above
Urine
Urinary Pb was detected
by mass spectrometry
Age at measurement:
50 yr
Not Reported
Vision: excellent, good, and
fair eyesight (self-reported)
were classified as good;
poor and very poor were
classified as poor
Hearing: good and little
trouble hearing (self-
reported) were classified as
good; lots of trouble and
deaf were classified as poor
Age, sex, ethnicity,
urine creatinine,
survey weighting
OR (95% Cl)b
Vision 1.15 (0.67-1.97)
Hearing 0.97 (0.63-1.51)
Balance 0.68 (0.51-0.91)
Balance: "During the past 12
mo, have you had dizziness,
difficulty with balance, or
difficulty with failing?"
3-468
-------
Referenc^and Study study Population Exposure Assessment Outcome Confounders Effect Estimates and
Ear ringing: "ears ringing,
roaring, or buzzing in the last
year"
Age at outcome:
50 yr
tKana et al. (2018)
Korea
2010-2013
Cross-sectional
KNHANES
n: 6409
Representative sample
of the entire Korean
population. Study
participants were at
least 20 yr old and
underwent pure-tone
audiometry and blood
Pb test.
Blood
Blood Pb was measured
using GFAAS and
classified into quartiles by
sex
Age at measurement:
20-87 yr (mean ± SE:
47.1 ± 0.3 yr)
Weighted mean ± SE
(Men):
|jg/dL;
|jg/dL;
|jg/dL;
pg/dL;
Weighted mean ± SE
(Women): Q1 = 1.12 ±
Q1 = 1.56 ± 0.01
Q2 = 2.22 ±0.01
Q3 = 2.82 ±0.01
Q4 = 4.22 ± 0.08
0.01 |jg/dL
0.01 |jg/dL
0.01 |jg/dL
0.03 |jg/dL
Q2 = 1.61 ±
Q3 = 2.11 ±
Q4 = 3.03 ±
Low-frequency hearing
impairment;
High-frequency hearing
impairment
Pure-tone audiometry was
performed on both ears at
0.5, 1, 2, 3, 4, and 6 kHz. A
binaural pure-tone average
threshold was used and two
binaural averages were
computed, one across 0.5, 1,
and 2 kHz and the other
across 3, 4, and 6 kHz to
determine the low- and high-
frequency thresholds.
Hearing impairment was
then determined according
to whether an average
threshold exceeded 25 dB in
the respective frequency
band.
Age at outcome:
20-87 yr (mean ± SE: 47.1 ±
0.3 yr)
Age, BMI, education,
smoking, alcohol
consumption,
exercise, diabetes
mellitus,
hypertension, noise
exposure
OR (95% Cl)b
Hearing Loss - Low
Frequency
Females
Q2
Q3
Q4
1.271 (0.726, 2.224)
1.308 (0.784, 2.183)
0.932 (0.541, 1.605)
Males
Q4
Q3
Q2
1.026 (0.813, 1.295)
1.028 (0.661, 1.598)
1.17 (0.772, 1.773)
Hearing Loss - High
Frequency
Females
Q2: 0.947 (0.608, 1.475)
Q3: 1.013 (0.698, 1.471)
Q4: 1.502 (1.027, 2.196)
Males
Q2: 1.368 (1.006, 1.86)
Q3: 1.402 (1.005, 1.955)
Q4:
1.629 (1.161, 2.286)
3-469
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tChoi and Park (2017)
Korea National Health and
Nutrition Examination Survey
(KNHANES), Korea
2010-2012
Cross-sectional
KNHANES
n: 5187 adults
Blood
Measured using Graphite
furnace atomic absorption
spectrometry
Age at Measurement: 20-
87 yr
90th: Adults: Geometric
mean (age-adjusted) 2.12
|jg/dL (95% CI: 2.08, 2.15)
Adolescents: Geometric
mean (age-adjusted) 1.26
|jg/dL (95% CI: 1.22, 1.30)
Hearing loss (>25dB) at
speech frequency;
Hearing loss (>25dB) at high
frequency
Pure-tone air conduction
hearing thresholds were
obtained for each ear at
frequencies of 0.5, 1, 2, 3, 4,
and 6 kHz over an intensity
range of-10 to 110 dB
-10 to 110 dB.
Age at outcome: 20-87 yr
Adjusted for age;
age squared; sex;
education; BMI;
current cigarette
smoking; current
diagnosis of
hypertension and
diabetes; and
occupational,
recreational, and
firearm noise
exposures
OR (95% Cl)b
Hearing Loss (>25 dB)
High-frequency PTA
Pb Quartile 2 (1.594-
2.146):
1.13 (0.83, 1.53)
Pb Quartile 3 (2.48-
2.822):
1.35 (1, 1.81)
Pb Quartile 4 (2.823-
26.507):
1.7 (1.25, 2.31)
Per doubling of Pb:
1.3 (1.08, 1.57)
Speech-Frequency PTA
Pb Quartile 2:
0.94 (0.65, 1.35)
Pb Quartile 3:
1.29 (0.92, 1.78)
Pb Quartile 4:
1.25 (0.87, 1.79)
Per doubling of Pb:
1.15 (0.94, 1.41)
tWanq et al. (2020)
Zhejiang Province
(Hangzhou, Jiangshan,
Tonglu, Jiaxing, Anji,
Jinyun), China
2016 to 2018
Case-control
n: 2016
Blood
Measured by graphite
furnace atomic absorption
spectrometry
Age at Measurement:
21-89 yr
Hearing loss
The devices utilized in this
research were an
audiometer (AT235,
Interacoustics AS, Assens,
Denmark) and standard
headphones (TDH-39,
Telephonies Corporation,
Farmingdale, USA)
Income, education,
hypertension,
diabetes,
hyperlipidemia, otitis
media, migraine,
anemia, smoking,
alcohol consumption,
daily fruit and
vegetable intake,
and workplace noise
exposure
OR (95% Cl)b Q1 Ref
Q2 1.135 (0.806, 1.599)
Q3 1.038 (0.731, 1.475)
Q4 1.016 (0.7, 1.475)
3-470
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Logarithmic-transformed
levels of Pb
Case group (1.58 ± 0.17
|jg/dL) and control group
(1.57 ± 0.16 Mg/dL)
Age at outcome: 21-89 yr
3-471
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tChoi etal. (2012)
NHANES, U.S.
1999-2004
Cross-sectional
NHANES
n: 3698
Blood
Simultaneous
multielement atomic
absorption spectrometer
(SIMAA 6000;
PerkinElmer, Norwalk, CT)
with Zeeman background
correction
Age at Measurement:
20-69 yr
Age-adjusted geometric
mean (95% CI) = 1.54
|jg/dL (1.49, 1.60)
Hearing threshold;
Hearing loss
Pure-tone air conduction
hearing thresholds were
obtained for both ears at
frequencies of 0.5-8 kHz
over an intensity range of-
10 to 120 dB.
Age at outcome: 20-69 yr
Age and age2, sex,
race/ethnicity [non-
Hispanic white
(reference), Mexican
American, non-
Hispanic Black,
other], education [<
high school
(reference), high
school, > high
school], BMI
(continuous),
ototoxic medication
use (yes/no),
cigarette smoking
[never smoker
(reference), < 20
pack-years, > 20
pack-years],
hypertension
(yes/no), type 2
diabetes (yes/no),
and either blood lead
or blood cadmium
(for the
corresponding
cadmium or lead
model), occupational
noise exposure
(0*NET score,
continuous),
nonoccupational
firearm noise
(yes/no) and any
recreational noise
(yes/no)
OR (95% Cl)b
Hearing Loss:
Quintile 2 (0.90-1.30
pg/dL)
1.08 (0.55, 2.12)
Quintile 3 (1.4-1.8 pg/dL)
1.1 (0.58, 2.05)
Quintile 4 (1.90-2.70
Mg/dL)
1.21 (0.67, 2.22)
Quintile 5 (2.80-54
MQ/dL)
1.36 (0.75, 2.48)
Per doubling of Pb
1.09 (0.95, 1.26)
3-472
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tYin etal. (2021)
Iran, Korea, China, United
States
Other
n: 234-7596 in 8
studies
Blood
Age at measurement:
3-87 yr
Hearing loss
All studies included
in the meta-analysis
controlled for age
and sex. Adjustment
for other potential
confounders varies
by studies, but
includes monthly
income, education
levels, smoking
status, BMI,
ethnicity, work
duration, ototoxic
medication, blood
lead, occupational
noise, loud noise,
and firearm noise,
and hypertension
and diabetes
OR (95% Cl)b
1.34 (1.18, 1.52)
tTu et al. (2021)
NHANES, U.S.
2011-2012
Cross-sectional
NHANES
n: 1503
Blood
Measured by plasma
mass spectrometry
Age at measurement:
20-69 yr
Median = 1.07 |jg/l
95th: 1.62 pg/l
Speech-frequency hearing
loss;
High-frequency hearing loss
For each ear, 0-5, 1, 2, 3, 4
and 6 kHz frequencies
were used for assessing
pure-tone air conduction
hearing thresholds
over a-10 to 110 dB
intensity ranges. The
average of four
audiometric frequencies
(0-5, 1, 2 and 4 kHz) was
used to identify
Age, sex, education,
marital status, BMI,
smoking, noise
exposure,
hypertension
and diabetes
OR (95% Cl)b
HFHL 1 98 (1-27, 3 10)
SFHL 1.46 (0.81, 2.64)
3-473
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
speech-frequency hearing
loss (SFHL), while the
average of three audiometric
frequencies (3, 4 and 6 kHz)
was used to identify
high-frequency hearing loss
(HFHL). SFHL or HFHL >25
dB in either ear was sued to
define hearing loss, based
on the WHO definition for
this condition
Age at outcome: 20-69 yr
tPaulsen et al. (2018)
Beaver Dam Offspring Study
Beaver Dam, Wisconsin,
U.S.
Baseline data collection
June 8, 2005, through
August 4, 2008 with two
follow-up examinations
occurred at 5-year intervals:
one was conducted between
July 12, 2010, and March 21,
2013, and the other between
July 1, 2015, and November
13, 2017
BOSS
n: 1983
Blood
Measured by Inductively
coupled plasma mass
spectrometry
Age at Measurement:
21-84 yr
Central tendency BLL: NR
Contrast sensitivity
impairment
Age, alcohol
consumption,
smoking, AMD,
cataract, plaque site,
VA impairment, and
sex
HR (95% CI)b
0.91 (0.696, 1.19)
Cohort
tFillion et al. (2013)
n: 228
Blood
Measured by Inductively
coupled plasma mass
spectrometry (ICP-MS)
Contrast sensitivity (cycles
per degree, cpd);
Acquired color vision loss
(color confusion index, CCI)
Age, sex, current
smoking (yes vs. no),
current drinking (yes
vs. no)
Beta (95% CI)b
Spatial frequency with
%EPA
3-474
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Lower Tapajos River Basin
State of Para
Brazil
Age at Measurement:
15-66 yr (median = 33.0
yr)
Age at outcome: 15-66 yr
1.5 cpd -1.32 (-4.30;
1.65)
3 cpd 2.06 (-2.87; 6.99)
6 cpd 0.60 (-6.04; 7.25)
12 cpd -13.33 (-23.28;
-3.49)
18 cpd -2.43 (-6.64;
1.79)
CCI 0.16 (-0.03; 0.33)
May to July 2006
Mean = 12.8 ± 8.4 pg/dL;
Median = 10.5 |jg/dL
Cross-sectional
BLL = blood lead level; CI = confidence interval; CIDI = Composite International Diagnostic Interview; EPA = eicosapentaenoic acid (; K-XRF = K-shell X-ray fluorescence; NAS =
Normative Aging Study; OR = odds ratio; Pb = lead; PTA = pure tone average; Q = quartile; RR= relative risk; SD = standard deviation; SE = standard error; WHO = World Health
Organization; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResult not standardized because data pertaining to the BLL distribution and/or base for the log-transformation were not reported.
tStudies published since the 2013 Integrated Science Assessment for Lead
3-475
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Table 3-16T Animal toxicological studies of Pb exposure and sensory organ function
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Jamesdaniel et al. (2018)
Mouse (C57BL/6)
Control (tap water), M, n =
6
2 mM, M, n = 6
PND 33 to
PND61
Oral,
drinking
water
PND 61:
10 |jg/L (1 pg/dL) for
Control
293 |jg/L (29.3 pg/dL) for
2 mM
PND 61: Auditory threshold (via BAEP)
Carlson et al. (2018)
Mouse (CBA/CaJ)
Control (deionized water),
M, n = 16
0.03 mM, M, n = 8
5 wkto 16
wk
Oral,
drinking
water
16 wk:
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure
Details
BLL as Reported (pg/dL)
Endpoints Examined
Liu etal. (2019)
Rat (Sprague Dawley)
PND 1 to
Oral,
PND 9: PND 93:
Sound-Azimuth Discrimination
Control (tap water), F, n =
PND 21
drinking
Training
12
water
0 pg/dL for Control
58 mg/L, F, n = 11
7.9 pg/dL for 58 mg/L
PND 21:
0 pg/dL for Control,
8.2 pg/dL for 58 mg/L
PND 40:
0 pg/dL for Control
0 pg/dL for 58 mg/L
BAEP = brainstem auditory evoked potentials; BLL = blood lead level; F = female; KNHANES = Korea National Health and Nutrition Examination Survey; M = male; Pb = lead; PND :
postnatal day; wk = week(s).
3-477
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Table 3-17E Epidemiologic studies of exposure to Pb and neurodegenerative disease in adults
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Wang et al. NAS
n: 358
NAS
US
Bone Pb
measurement
(1991-1999),
Mini-Mental State
Examination
(MMSE) twice
(1993-1998 and
1995-2000)
Tibia and patella
Measured by K-XRF
Age at measurement 21-
81 yr
Median: 19 and 23 |jg/g
for tibia and patella
Cognitive decline Adjusted for age, years of
cognitive assessment battery was education, nonsmoker,
the MMSE, a global examination f°rmer smoker, pack-years,
of cognitive function that
assesses orientation, immediate
and short-term recall, verbal and
written skills, and attention and
ability to follow commands
Age at outcome: 21-81 yr
nondrinker, alcohol
consumption, English as first
language, computer
experience, and diabetes
Change in MMSE score
per IQR (15 |jg/g)
increase in tibia Pb by
class of HFE genotype3
Wild-type -0.02 (-0.10
to 0.07)
One HFE variant allele
-0.14 (-0.33 to 0.04)
Two HFE variant alleles
-0.63 (-1.04 to-0.21)
Cross-sectional
3-478
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Weisskopf et al. NAS
12004} n: 466
Normative Aging
Study, U.S.
1991 and 2002
Cross-sectional
Tibia, patella, and blood
Bone Pb measured by
ABIOMED K-XRF
instrument and blood Pb
measured by Zeeman
background-corrected
flameless atomic
absorption (graphite
furnace)
Age at measurement 21-
81 yr
Cognitive decline
cognitive assessment battery was
the MMSE, a global examination
of cognitive function that
assesses orientation, immediate
and short-term recall, verbal and
written skills, and attention and
ability to follow commands
-0.25 (-0.45, -0.05)
Age at first MMSE test, Difference in change in
alcohol intake, and days MMSE score per IQR
between the two MMSE tests increase in Pba
as continuous variables, as
well as education (<12 yr, 12
yr, 13-15 yr, >16 yr), smoking
status (never, former,
current), computer
experience (yes/no), and
English as a first language
(yes/no)
Median |jg/g (interquartile
range)
Patella 27 (19, 40) pg/g
Tibia 21 (15, 29) pg/g
Blood 5 (3, 7) pg/dL
3-479
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
(Wright et al.. NAS
2003) n: 1033
Normative Aging
Study, U.S.
1991-1997
Cross-sectional
Tibia, patella, and blood
Bone Pb measured by
ABIOMED K-XRF
instrument and blood Pb
measured by Zeeman
background-corrected
flameless atomic
absorption (graphite
furnace)
Cognitive decline Age, alcohol intake, and
cognitive assessment battery was education history
the MMSE, a global examination
of cognitive function that
assesses orientation, immediate
and short-term recall, verbal and
written skills, and attention and
ability to follow commands
OR (95% Cl)a MMSE
<24
Tibia 1.02 (1.00,1.04)
Patella 1.02 (1.00,
1.03)
Age at measurement 21-
81 yr
Mean (SD)
Patella 29.5 (21.2) pg/g
Tibia 22.4 (15.3) pg/g
Blood 4.5 (2.5) pg/dL
3-480
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
(Weuve et al.. NAS
2006) n: 1171
Normative Aging
Study, U.S.
1991 and 2002
Cross-sectional
Tibia, patella, and blood
Measured by graphite
furnace atomic
absorption with Zeeman
background correction
Age at measurement 21-
81 yr
Median (and first and
third quartiles) of tibia
and patella were 19 (13,
28) and 27 (18, 39) pg/g
Blood 5.2 (<1-28) pg/dl
ALAD genotype modifications on
cognition
cognitive assessment battery was
the MMSE, a global examination
of cognitive function that
assesses orientation, immediate
and short-term recall, verbal and
written skills, and attention and
ability to follow commands
Age at cognitive assessment
and age-squared, years of
education (<8, 9-11, 12, 13-
15, 16, >17 yr), computer
experience (an additional
measure of socioeconomic
status), and length of time
between the lead and
cognitive assessments, were
smoking status (current v
past or never), alcohol
consumption (none, 0.1-4.9
g/day, 5.0-9.9 g/day, >10
g/day, or missing), calorie
adjusted calcium intake (in
tertiles), regular energy
expenditure on leisure time
physical activity (in tertiles),
and diabetes (physician
diagnosed or fasting blood
glucose >126 mg/dl)
Mean difference in
MMSE score per IQR
increase in Pb (95%
Cl)a
Tibia
Among ALAD-2
carriers -0.16 (-0.58 to
0.27)
Among ALAD wildtypes
-0.05 (-0.21 to 0.12)
Patella
Among ALAD-2
carriers -0.26 (-0.64 to
0.12)
Among ALAD wildtypes
-0.07 (-0.23 to 0.09)
Blood
Among ALAD-2
carriers -0.26 (-0.54 to
0.01)
Among ALAD wildtypes
-0.04 (-0.16 to 0.07)
(Nordbera et al.. Kungsholmen project
2000) n: 762
Stockholm
1994-1996
Cross-sectional
Blood
Measured using Graphite
furnace atomic
absorption spectrometry
Age at measurement:
75+ (mean age of 88.4
yr)
Mental Performance
MMSE
Age and BMI
No association was
reported (quantitative
estimate NR)
Mean (SD) 3.7 (2.3) pg/dl
3-481
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tFarooqui et al. Participants selected from Bone
(2017)
Boston, MA,
United States
1993-2007
Cohort
Changes in cognition
cohort study (Veterans
Affairs NAS); healthy men
aged 51-98 yr
n: 741 subjects in MMSE
and 715 in Global cognition
Age at first cognitive test,
Patella (trabecular bone)
and tibia (cortical bone)
bone Pb was measured
using K-XRF
spectroscopy
in 1993
Patella mean (SD) 30.6 ±
19.44 |jg/g, and tibia
mean (SD)21.6± 13.33
pg/g
past education level, baseline
Cognition was assessed using smoking status and alcohol
the MMSE, NES2, CERAD and intake.
WAIS-R during 3-5 visits over the
period of 15 yr of follow-up.
HR (95% Cl)b
Cognition
MMSE < 25
Tibia
1.05 (0.82, 1.35)
Patella
1.21 (0.99, 1.49)
Beta (95% Cl)b
Global Cognition
(Summary score of
NES2, CERAD and
WAIS-R)
Tibia
-0.206 (-0.453, 0.089)
Patella
-0.25 (-0.518, 0.019)
Cognition
MMSE
Tibia
-0.077 (-0.206, 0.052)
Patella
-0.128 (-0.251,
-0.0004)
3-482
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tYanq et al.
(2018)
Multicity
(Taichung city,
Changhua, and
Nantou County),
Taiwan
Feb 2015-Oct
2016
Case-control
Participants were recruited
from the China Medical
University Hospital. Cases
recruited the Department of
Neurology, and controls
from the
Department of Family
Medicine who were
receiving general health
check-up; aged >50 yr.
n: Full sample: 434 (170 AD
and 264 controls);
Propensity-score-matched
sample: 84 AD and 84
controls.
Blood assessed for heavy AD risk
Age, gender, education,
metals
Blood samples tested for
heavy metals (Pb, Cd,
Se, Hg). Blood Pb
measured through ICP-
MS
Blood Pb cone: Full
samples: AD: 2.50 ± 1.35
|jg/dL, Controls: 2.36 ±
1.02 |jg/dL
Propensity matched
samples: AD: 2.58 ± 1.35
|jg/dL, Controls: 2.50 ±
1.18 |jg/dL
exercise habits
Physician AD diagnosis based on hypertension, diabetes
definition of the Diagnostic and
Statistical Manual Fourth Edition
Criteria; MMSE test of cognitive
function
cardiovascular diseases,
depression, anxiety
OR (95% Cl)b
Full population
Total: 1.05 (0.86-1.28)
Tertile 2 vs. 1: 1.00
(0.56-1.79)
Tertile 3 vs. 1: 0.87
(0.49-1.55)
Propensity score-
matched population
Total: 1.06 (0.83-1.35)
Tertile 2 vs. 1: 1.16
(0.55-2.47)
Tertile 3 vs. 1: 1.12
(0.53-2.39)
tHorton et al.
(2019)
Nationwide,
United States
1999-2008
followed till 2014
Cohort
Participants selected from Blood assessed for Pb AD mortality
the five NHANES cycles and
included 1999 to 2008 who
were followed till 2014 for
death; aged >60 yr.
n: 8,080 subjects
Blood samples collected
during the NHANES
mobile examination
center visit assessed for
Pb using ICP-DRC-MS.
Blood Pb cone: Geo
mean and 95% CI: 2.1
(2.02, 2.11) |jg/dL
The identification of AD mortality
is based on the immediate cause
of death in the National Death
Index record. Cause of death was
coded according to the ICD-10,
revision 10; G30 was used to
indicate AD.
Age, sex, poverty status,
race/ethnicity, and smoking
status, and competing risks
for AD mortality.
HRR (95% CI)b
0.3 |jg/dL:
ref
0.5 |jg/dL:
1.1 (0.89, 1.3)
1 |jg/dL:
1.2 (0.77, 1.8)
1.5 |jg/dL:
1.2 (0.7, 2.1)
2 |jg/dL:
1.3 (0.66. 3)
3 |jg/dL:
1.3 (0.6, 3.0)
5 |jg/dL:
1.4 (0.54, 3.8)
3-483
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
(Vinceti et al.. n: 15 cases and 36 controls Blood
1997)
ALS measured by ALS severity Patients and controls
Correlation coefficient
scale
Santa Mafia
Nuova Hospital
in Reggio Emilia,
northern Italy
December 31st,
1994
Case-control
Age at measurement:
(mean ±SD)
Patients 65.9 ± 14.0 yr
Controls 64.4 ± 12.9
Mean (SD)
Controls 108.3 ± 44.4
pg/i
Patients 127.1 ± 67.8 pg/l
matched on year of birth and ALSSS (p-value)b
gender, confounders NR Tota| -0.440 (0.101)
(Kamel et al.
2002)
New England
and U.S.
1993-1996
Case-control
n: 109 cases and 256
controls
Blood and bone
Blood lead was
measured using graphite
furnace atomic
absorption spectrometry
ALS
A board-certified neurologist (T.
L. M. or J. M. S.) evaluated
potential cases. Diagnosis of ALS
was based on criteria published
by the World Federation of
Cases and controls matched
on age, sex, and region
OR (95% Cl)b
Blood 1.9 (1.4, 2.6)
Tibia 2.3 (0.4, 14.5)
Patella 3.6 (0.6, 20.6)
Bone lead was measured Neurology,
using in vivo K-XRF
Age at measurement:
30-80 yr
Mean (SE)
Blood |jg/dl
Cases 5.2 (0.4)
Controls 3.4 (0.4)
Patella |jg/g
Cases 20.5(2.1)
Controls 16.7 (2.0)
Tibia |jg/g
Cases 14.9 (1.6)
Controls 11.1 (1.6)
3-484
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Kamel et al. n: 110
(2008)
New England
and
U.S.
Enrollment: 1993
- 1996; follow-up
through
December 31,
2003
Cohort
Bone and blood
Bone Pb measured in the
tibia and patella using K-
XRF; blood Pb measured
using atomic absorption
spectrometry
Neurodegenerative Disease ¦
ALS
Age at Measurement:
Median (range
79) years
Amyotrophic lateral sclerosis
(ALS) was diagnosed by board-
certified neurologists and based
on the World Federation of
Neurology El Escorial criteria;
related symptoms were
documented from interviews.
60 (30- Cause of death was identified by
the National Death Index (NDI).
Cox proportional hazard
analyses adjusted for age,
sex, and ever smoked,
except for sex-stratified
models, which included age
and ever smoked.
HR (95% Clf
Diagnosis to death 0.9
(0.8 to 1.0)
Symptoms to death 0.9
(0.8 to 1.0)
Blood Pb median = 4
|jg/dL; patella Pb median
= 15 |jg/g; tibia Pb mean
= 13 |jg/g
Max: Blood Pb max = 14
|jg/dL; patella Pb max =
107 pg/g; tibia Pb max =
61 pg/g
(Fang et al..
2010)
U.S.
2003-2007
Case-control
n: 184 cases and 194
controls
Blood
Pb measured by
inductively coupled
plasma mass
spectrometry
Age at measurement:
mean (SD)
Cases 63.3 (34-83)
Controls 63.4 (34-84)
Pb Mean (SD)
3.4 pg/dL (2.5)
ALS
Neurologists with expertise in
ALS reviewed medical records to
determine motor neuron disease
diagnosis in accordance with the
original El Escorial Criteria,
including ALS (International
Classification of Diseases, Ninth
Revision (ICD-9) code 335.20),
progressive muscular atrophy
(ICD-9 code 335.21), progressive
bulbar palsy (ICD-9 code 335.22),
pseudobulbar palsy (ICD-9 code
335.23), primary lateral sclerosis
(ICD-9 code 335.24), and other
motor neuron diseases (ICD-9
code 335.29).
Age
OR (95% Cl)b
Overall 1.9 (1.3, 2.7)
3-485
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tFanq et al.
(2017)
Nationwide,
United States
2007-2013
Cohort
Veterans with ALS in the
U.S. National Registry of
Veterans, and other
veterans with ALS not
treated within the Veterans
Affairs healthcare system,
with ALS from April 2003 to
September 2007 and
followed till the date of death
or July 25, 2013; non-
Hispanic Caucasian men
aged 34-83 yr
n: 145 U.S. Veterans with
ALS, who were male,
diagnosed with ALS by
neurologist.
Blood assessed for Pb ALS survival
Whole blood collected
during Jan-Sep 2007
assessed for Pb using
ICP-MS.
Biomarkers for bone
formation measured in
plasma. Bone formation
was measured using
procollagen type I N-
terminal propeptide and
bone resorption
measured using C-
terminal collagen
crosslinks.
Blood Pb cone: 2.35 ±
1.28 |jg/dL
Age at diagnosis, diagnostic
certainty, site of onset,
Trained neurologist with expertise diagnostic delay and revised
in ALS assigned diagnoses using ALS Functional Rating Scale
an algorithm based on the Score
revised El Escorial Criteria
HR (95% Clf
1.234 (1.021, 1.49)
tPeters et al.
(2020)
EPIC
Multi-center,
Europe
1993-1999
Nested case-
control
N = 107 cases identified
after 8 yr of follow-up
3 controls per case
Pb concentration in
erythrocytes analyzed
using ICP-MS
ALS: Motor neuron disease
(ICD10 G12.2) as underlying
cause of death
Matched by age at OR (95%CI)
recruitment, sex, study center Reference: <56.8 ng/g
>56.8-<89.0: 1.83
(0.99, 3.35)
>89.0: 1.89 (0.97, 3.67)
tVinceti et al.
(2017)
Emilia-Romagna;
Italy
May 1998—April
2011
Case-control
Cases were ALS patients
and controls were selected
from hospital-admission of
no ALS; mean age cases:
52 yr
n: 76 (38 ALS cases and 38
controls)
Cerebral spinal fluid
assessed for heavy
metals
CSF evaluated for heavy
metals (Pb, Cd, Hg) using
ICP-MS
Median Pb cone: Cases:
155 ng/L, Controls: 132
ng/L
ALS
Probable ALS diagnosis using the
revised El Escorial Criteria.
Age, sex, and total selenium
In the highest fertile of
exposure, OR (95%
Cl)b 1.39 (0.48, 4.25)
3-486
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tAndrew et al.
(2022)
Nationwide,
United States
2013-2019
Case-control
Participants are from the
healthcare claims dataset
from Symphony Health with
ALS diagnosis after 6 mo
enrollment in the database
prior to the first ALS ICD
code. Controls are
individuals similar to ALS
cases based on age, sex,
and length of database
history with min of 6 mo in
database; cases and
controls age 18-80 yr (63%
were 55-75 yr)
n: Cases: 26,199 and
controls: 78,597
268 Airborne
contaminants
Airborne exposure to Pb
and other contaminants
assessed from U.S.
EPA's NEI database for
2008 to estimate
exposure prior to ALS
onset. Data was used to
estimate residential
exposure at the zip3
locations of the ALS
patients and controls.
ALS
ALS based on the healthcare
claims
Family income, race, age,
and sex
OR (95% Cl)b
Discovery and
Validation Cohorts
1.39 (95% CI 0.48-
4.25)
New
Hampshire/Vermont
[Q1+Q2: <1.37 tons
(Ref)
Q3: 1.37-26.1
Q4: >26.1]
5-year Exposure
History
Q3: 1.79 (1.32, 2.43)
Q4: 1.11 (0.8, 1.55)
10-year Exposure
History
Q3: 2.42 (1.76, 3.33)
Q4: 2.03 (1.46, 2.8)
15-year Exposure
History
Q3: 1.83 (1.34, 2.52)
Q4: 1.73 (1.26, 2.38)
Ohio
[Q1+Q2: <14.7 tons
(Ref) Q3: 14.7-50.8
Q4: >50.8]
3-487
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
5-year Exposure
History
Q3: 0.48 (0.37, 0.61)
Q4: 0.39 (0.3, 0.51)
10-year Exposure
History
Q3: 1.05 (0.83, 1.33)
Q4: 1.6 (1.28, 1.98)
15-year Exposure
History
Q3: 0.94 (0.74, 1.18)
Q4: 1.07 (0.86, 1.34)
3-488
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tPaul et al.
(2021)
Australia and
New Zealand,
and Central
California
Case-control
Participants for this study
comes from two publicly
available PD studies: SGPD
consortium of three studies
across Australia and New
Zealand with cases and
controls. PEG a population-
based study from three
agricultural counties of
Central California with cases
and controls.
n: SGPD cohort: 959 cases
and 930 controls; PEG
cohort: 569 cases and 238
controls
Epigenetic biomarkers for PD
cumulative Pb exposure
(tibia and patella), i.e.,
DNAm Pb
Epigenetic biosensors
identified with site-by-sire
analysis and combined
via machine learning
algorithm on K-XRF in
vivo measures of bone
Pb. To determine Pb
biomarker level in two
cohorts, the regression
coefficients were
extracted from NAS and
applied to the DNAm beta
matrices.
Age (DNAm Age in SGPD),
sex, ancestry (PEG only),
smoker (PEG only), blood
cell composition, and mean
Meth By Sample
OR (95% Cl)b
Tibia SPGD
1.54 (1.22, 1.95)
Patella SPGD
0.70 (0.53, 0.93)
Tibia PEG
1.52 (1.25, 1.86)
DNAm tibia Pb: SGPD
cohort: cases: 3.41 ± 0.4,
controls: 3.48 ± 0.4; PEG
cohort: cases: 3.06 ± 0.4,
controls: 3.03 ± 0.3
tJietal. (2015)
Participants selected are
Blood, Bone assessed for Tremor
Age, age squared, alcohol
OR (95% Cl)b
subgroup of participants
Pb
consumption, smoking status,
Blood
Boston, MA,
United States
from cohort study (Veterans
Affairs NAS); healthy men
Blood samples tested for
education level
Quintile 2 vs. 1
aged 50-98 yr.
Pb concentration using
0.09 (-0.10, 0.29)
NAS
Zeeman background-
Quintile 3 vs. 1
n: 807
corrected flameless
0.06 (-0.16, 0.28)
atomic absorption
Cohort
graphite furnace. Bone
Quintile 4 vs. 1
Pb concentration
measured with K-XRF at
0.07 (-0.13, 0.27)
Quintile 5 vs. 1
both the tibia and the
0.07 (-0.16, 0.30)
Patella
patella starting in 1991.
Blood Pb concentration:
Quintile 2 vs. 1
5.01 ± 2.72 |jg/dL
3-489
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
Tibia Pb cone: 21.23 ±
13.29 |jg/g (); patella Pb Tremor score was created based
cone: 27.98 ± 18.38 |jg/g) on an approach using hand-
drawing samples that were
derived from figure copying
testing performed as part of
larger cognitive test battery
(CERAD, MMSE, and VMI)
assessed over a mean follow-up
of 8.0 ± 3.2 yr after bone Pb
measurement.
ALAD genotype was determined
by amplifications of 0.5 |jL of
whole blood using two sets of
primers specific for a portion of
the ALAD gene.
-0.07 (-0.30, 0.15)
Quintile 3 vs. 1
-0.14 (-0.37, 0.09)
Quintile 4 vs. 1
-0.22 (-0.45, 0.01)
Quintile 5 vs. 1
-0.02 (-0.27, 0.22)
Tibia
Quintile 2 vs. 1
0.03 (-0.20, 0.26)
Quintile 3 vs. 1
0.03 (-0.20, 0.26)
Quintile 4 vs. 1
0.13 (-0.11, 0.36)
Quintile 5 vs. 1
-0.07 (-0.32, 0.17)
3-490
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tKhalil et al.
(2014)
Pittsburgh, PA
United States
2007-2009
Cross-sectional
MrOS
n: 445
Non-Hispanic Caucasian
men (community dwelling
non-institutionalized) at least
65 yr of age enrolled in
MrOS at the University of
Pittsburgh clinic. Eligibility
criteria included the ability to
walk unaided and without
bilateral hip replacements.
Blood Grip strength (kg);
Leg extension power (watts);
Blood Pb measured using Walking speed (m/s);
AAS Narrow-walk pace (m/s);
Age at measurement: Use arms to stand up (yes/no)
Mean = 79.5 ± 5 yr
Grip strength was measured on a
Mean = 2.25 |jg/dL; SD = Jamar dynamometer;
1.20 |jg/dL; Median = 2 Leg extension power was
Age, education, smoking,
alcohol consumption, BMI
Beta (95% Clf
Leg extension power
-0.03 (-1.97, 2.03)
Ability to stand from a
chair without using their
arms 0.97 (0.88, 1.07)
|jg/dL
Max: 10 pg/dL
measured with the Nottingham
power rig;
Walking speed was assessed on
a standard 6-m walking course;
Narrow-walk pace (an indirect
measure of dynamic balance)
was assess while keeping each
foot within a 20-centimeter wide
lane on the 6-m walking course;
Stand from a chair without using
the arms was measured as
yes/no.
Age at outcome:
65 yr
3-491
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tShiue (2013)
United States
2003-2004
Cross-sectional
NHANES
n: 712 (vision); 732
(hearing); 669 (balance)
Urine
Urinary Pb was detected
by mass spectrometry
NHANES age 50 and above Age at measurement:
50 yr
Not Reported
Vision;
Hearing;
Balance;
Ear ringing
Vision: excellent, good, and fair
eyesight (self-reported) were
classified as good; poor and very
poor were classified as poor
Hearing: good and little trouble
hearing (self-reported) were
classified as good; lots of trouble
and deaf were classified as poor
Balance: "During the past 12 mo,
have you had dizziness, difficulty
with balance, or difficulty with
failing?"
Ear ringing: "ears ringing, roaring,
or buzzing in the last year"
Age at outcome:
50 yr
Age, sex, ethnicity, urine
creatinine, survey weighting
OR (95% Cl)b
Vision 1.15 (0.67-1.97)
Hearing 0.97 (0.63-
1.51)
Balance 0.68 (0.51 —
0.91)
3-492
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tCasiens et al.
(2018)
Ruhr area, a
German
industrial region
with a high
volume of steel
production
Germany
Baseline
recruitment
2000-2003;
Follow-up 2011-
2014
Cohort
HNRS
n: 1188
Men from the Heinz Nixdorf
Recall Study. Recruitment
details not provided.
Blood
Blood Pb was measured
in aliquots of whole blood
archived at baseline and
at follow-up using ICP-
MS
Age at measurement:
Median = 58 yr at
baseline (range 45-75 yr)
and 68 yr (range 55-86)
yr at follow-up
Median = 3.29 (IQR
2.55-4.32) |jg/dL at
baseline; 2.59 (IQR 1.99—
3.39) |jg/dL at follow-up
Max: 67.73 |jg/dL at
baseline; 39.68 |jg/dL at
follow-up
Odor identification;
Tapping hits;
Aiming errors;
Line tracing errors;
Steadiness errors
Odor identification: Sniffin sticks
odor identification test of 12
odors, participants classified as
normosmic if >9 odors identified,
hyposmic if 7-9 odors identified,
and functionally anosmic if <7
odors identified
Tapping hits: tapping a stylus
within 32 s as fast as possible;
- hits <10th percentile were
considered as substantially
impaired manual dexterity
Aiming errors: 20 small plates
with a diameter of 5 mm standing
in a line (distance 4 mm) had to
be touched with a stylus as fast
as possible; errors >90th
percentiles were considered as
substantially impaired manual
dexterity
Line tracing errors: drawing a
stylus through a curvy course of a
groove without touching side
walls or bottom; errors >90th
percentiles were considered as
substantially impaired manual
dexterity
Steadiness errors: maintain a
precise arm-hand position by
holding a stylus for 32 s in a 5.8
mm hole without touching sides
or bottom; errors >90th
percentiles were considered as
substantially impaired manual
dexterity
Age at outcome:
55-86 yr
Occupational qualification,
age, smoking status, alcohol
consumption, total test time
OR (95% Cl)b
Group 1: <5 |jg/dL (Ref)
Group 2: 5-<9 |jg/dL
Group 3: >9 |jg/dL
Odor identification
baseline
G2: 0.91 (0.65, 1.28)
G3: 1.96 (0.94, 4.11)
follow-up
G2: 1.04 (0.55, 1.94)
G3: 1.57 (0.47, 5.19)
Motor Performance
Series
Steadiness Errors
baseline
G2: 0.99 (0.62, 1.59)
G3: 1.36 (0.5, 3.66)
follow-up
G2: 1.16 (0.5, 2.69)
G3: 1.75 (0.41, 7.58)
Line tracing errors
baseline
G2: 1.09 (0.68, 1.76)
G3: 0.93 (0.32, 2.74)
follow-up
G2: 1.01 (0.41, 2.48)
G3: 0.59 (0.08, 4.11)
Aiming errors
baseline
G2: 1.07 (0.75, 1.53)
G3: 0.56 (0.22, 1.42)
3-493
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
follow-up
G2: 1.35 (0.73, 2.51)
G3: 0.42 (0.09, 2.08)
Tapping Hits
baseline
G2: 0.87 (0.53, 1.44)
G3: 1.35 (0.49, 3.7)
follow-up
G2: 2.63 (1.26, 5.49)
G3: 0.8 (0.14, 4.59)
3-494
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tJietal. (2013)
United States
1999-2002
Cross-sectional
NHANES
n: 3,593 (1,798 women;
1,795 men)
NHANES data from the
1999-2000 and 2001-2002
surveys, participants were
50 yr of age
Blood
Blood Pb was measured
using AAS
Age at measurement:
50-85 yr (Median = 61.2
yr)
Mean ± SD: Women =
2.17 ± 0.04 |jg/dL; Men =
3.18 ± 0.08 |jg/dL
("there's a slight
discrepancy in the SDs in
Table 1 vs. text on p.
712)
Median: Women = 1.72
|jg/dL; Men = 2.41 |jg/ dL
Walking speed (ft/sec)
Time to walk 20 ft (at usual
walking pace)
Age at outcome:
50-85 yr (median = 61.2 yr)
Model 4 (fully adjusted): age,
education, ethnicity, height,
waist circumference, alcohol,
smoking, physical activity,
arthritis, diabetes, heart
condition, hypertension,
homocysteine, C-reactive
protein
Beta (95% CI)
Walking Speed-Men
4.4 to <54.0
-0.029 (-0.155, 0.097)
Walking Speed-Women
3.0 to <53.0
-0.114 (-0.191,
-0.038)
Walking Speed-Men
3.1 to < equal to 4.3
0.082 (-0.012, 0.176)
Walking Speed-Women
2.2 to <2.9
-0.104 (-0.187,
-0.021)
Walking Speed-Men
2.4 to <3.0
-0.17 (-0.26, -0.08)
Walking Speed-Women
1.7 to <2.1
-0.024 (-0.118,
-0.063)
Walking Speed-Men
1.8 to <2.3
0.057 (-0.051, 0.165)
Walking Speed-Women
1.3 to <1.6
-0.027 (-0.055, 0)
3-495
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tMin etal. (2012) NHANES
n: 5574
United States
1999-2004
Cross-sectional
Adults who participated in
the NHANES Balance
Component and had blood
Pb and Cd measurements
and data for all covariate
variables
Blood
Blood Pb was measured
using a multielement AAS
with Zeeman background
correction.
Age at measurement:
40 yr
Weighted mean
(participant without
balance dysfunction):
2.09 |jg/dL (95% CI: 2.01,
2.18); Weighted mean
(participants with balance
dysfunction): 2.39 |jg/dL
(95% CI: 2.29, 2.49)
Max: 48 pg/dL
Balance dysfunction
Balance dysfunction was
evaluated by the Romberg Test
of Standing Balance on Firm and
Compliant Support Surfaces,
which measured the participant's
ability to maintain balance under
four test conditions: Test 1)
maintain balance while standing
(with feet together and arms
folded across the waist, holding
each elbow with the opposite
hand) for 15 sec.; Test 2)
maintain balance while standing
for 15 sec with eyes closed so
that only vestibular and
proprioceptive (i.e., leg muscle
position sense) information is
available; Test 3) maintain
balance while standing for 30 sec
on a foam-padded surface, which
reduces proprioceptive input but
does not affect visual or
vestibular input; Test 4) maintain
balance while standing for 30 sec
on a foam-padded surface with
eyes closed, so that input is
available from the vestibular
system only. Each condition was
scored on a pass or fail basis.
The time to failure (i.e., loss of
balance) was also recorded for
test condition 4, with those who
passed the test assigned the
maximum value of 30 sec.
Age, sex, race/ethnicity,
education, pack-years of
smoking, alcohol
consumption, histories of
stroke and diabetes, intakes
of Ca2+ and iron
OR (95% CI)
Balance Dysfunction
(Quintile 5 [3.3—48
Mg/dL])
33.334 (1.939,
573.157)
Balance Dysfunction
(Quintile 4 [2.3-3.2
MQ/dL])
5.234 (0.59, 46.429)
Balance Dysfunction
(Quintile 3 [1.8-2.2
MQ/dL])
0.665 (0.05, 8.783)
Balance Dysfunction
(Quintile 2 [1.3-1.7
Mg/dL])
3.707 (0.544, 25.282)
Age at outcome:
40 yr
3-496
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs
tGrashow et al.
(2013)
Greater Boston
area, MA, United
States
Grooved
pegboard May
2005- December
2009
Neuroskill July
2004 and
November 2007
Cohort
Normative Aging Study
n: 362 for grooved pegboard
test; 328 for the Neuroskill
test
Elderly, majority Caucasian
men originally recruited from
the greater Boston,
Massachusetts area in the
1960s
Bone
Bone Pb was measured
at the patella and the
midtibial shaft using an
ABIOMED K-XRF
instrument. "Tibia and
patella bone Pb
concentrations reflect
cumulative Pb exposure
over different time
windows: patella Pb
reflects exposure over
the last decade, while
tibia Pb half-life is on the
order of decades"
Age at measurement:
NR
Mean patella Pb = 25.0
mg/g bone (SD = 20.7);
Mean tibia Pb = 19.2
mg/g bone (SD = 14.6)
Grooved pegboard (completion
time, seconds);
Neuroskill (Signature score, %);
Neuroskill (Im pattern score, %)
Grooved pegboard test: Subjects
were asked to insert the metal
pegs into each of the 25 holes in
sequence as quickly as possible
with their dominant hand without
practice trials;
Neuroskill tests (signature score,
%): Subjects were asked to
provide five samples of their
signature in succession, written in
their natural manner;
Neuroskill tests (Im pattern score,
%): Subjects were asked to
provide five samples of a series
of cursive Ims (Im pattern) using
the instrumented pen
Age at outcome:
Mean age = 69.1 yr (SD = 7.2)
Age, smoking, education, Beta (95% Cl)a
computer experience, income Neuroskill-lm pattern
score
Patella 0.45 (0.178,
0.723)
Tibia 0.847 (0.163,
1.53)
Neuroskill-Signature
Score
Patella 0.08 (-0.47,
0.63)
Tibia -0.293 (-1.063,
0.477)
Grooved pegboard-
dominant hand
completion time
Patella 1.965 (0.553,
3.378)
Tibia 3.107 (1.157,
5.057)
AAS = atomic absorption spectrometry; BLL = blood lead level; Cd = cadmium; CERAD = Consortium to Establish a Registry for Alzheimer's Disease; CI = confidence interval; CSF =
cerebrospinal fluid; EE = effect estimate(s); HNRS = Heinz Nixdorf Recall Study; HRR = hazard rate ratio; K-XRF = K-shell X-ray fluorescence; MMSE = Mini Mental State
Examination; mo = month(s); MrOS = Osteoporotic Fractures in Men Study; NAS = Normative Aging Study; NEI = National Emissions Inventory; NHANES = National Health and
Nutrition Examination Survey; OR = odds ratio; Pb = lead; SD = standard deviation; sec = second(s); VMI = visual-motor integration; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bResult not standardized because data pertaining to the BLL distribution and/or base for the log-transformation were not reported.
tStudies published since the 2013 Integrated Science Assessment for Lead.
3-497
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Table 3-17T Animal toxicological studies of Pb exposure and neurodegeneration
Study
Species (Stock/Strain), n, Timing of Exposure
Sex Exposure Details
BLL as Reported (pg/dL)
Endpoints Examined
Zhou etal. (2018)
Rat (Sprague Dawley)
Control (distilled water), M, n
= 10
0.5% solution, M, n = 10
1.0% solution, M, n = 10
2.0% solution, M, n = 10
PND 24 to PND 52 Oral, drinking
water
PND 52:
13.3 |jg/L (1.3 pg/dL) for Control
148.9 |jg/L (14.9 pg/dL) for 0.5%
solution
231.3 |jg/L (23.1 pg/dL) for 1.0%
solution
293.4 |jg/L (29.3 pg/dL) for 2.0%
solution
PND 24, 31, 38, 45,
52: Amyloid protein
expression, Brain
Cholesterol,
Expression of BACE1
and APP
Li etal. (2016c)
Mouse (Kunming)
Control (distilled water), M/F,
n = 10
0.1% solution (mass fraction),
M/F. n = 10
0.2% solution (mass fraction),
M/F, n = 10
0.5% solution (mass fraction),
M/F, n = 10
GD to PND 21
Oral, lactation
In utero
PND 21:
10.62 pg/L (1.1 pg/dL) for Control
40.71 pg/L (4.1 pg/dL) for 0.1%
solution
81.77 pg/L (8.2 pg/dL) for 0.2%
solution
103.36 pg/L (10.3 pg/dL) for 0.5%
solution
PND 21: Amyloid
protein expression
Gu etal. (2012)
Mouse (Tg-SwDI)
4-8 wk to 10-14 wk
Oral, gavage
10-14 wk:
10-14 wk: Beta-
Control (Na-acetate water),
amyloid and APP
NR, n = 4-7
1.83 pg/dL for Control
expression
50 mg/kg, NR, n = 4-7
29.5 pg/dL for 50 mg/kg
Wu et al. (2020b)
Mouse (C57BL/6)
Control (distilled deionized
water), M, n = 7-10
0.2% solution, M, n = 7-10
4 wk to 4 mo
Oral, drinking
water
16 mo:
66.4 pg/L (6.6 pg/dL) for Control
278.9 pg/L (27.9 pg/dL) for 0.2%
solution
16 mo: Expression of
BACE1 and APP,
Phosphorylated tau
expression
3-498
-------
Study
Species (Stock/Strain), n, Timing of Exposure
Sex Exposure Details
BLL as Reported (pg/dL)
Endpoints Examined
Sun et al. (2014)
Rat (Sprague Dawley)
Control (tap water), NR, n :
20
580 ppm, NR, n = 20
NR (230-260 g) - 3 Oral, drinking
mo of treatment water
After 3 mo treatment:
3.0 |jg/L (0.3 pg/dL) for Control
56.8 |jg/L (5.7 pg/dL) for 580 ppm
After 3 mo treatment:
Immunohistochemistry
of APP
Gassowska et al. (2016b)
Rat (Wistar)
GD Oto PND21
Oral, lactation
PND 28:
PND 28: Tau protein
Control (tap water), M/F, n = 8
In utero
expression and
0.93 pg/dL for Control
phosphorylation
0.1% solution, M/F, n = 8
6.86 pg/dL for 0.1% solution
Rahman et al. (2012b)
Rat (Wistar)
PND 1 to PND 30
Oral, drinking
PND 21:
PND 21, 30: Tau
Control (tap water), M/F, n =
water
protein expression
6-10
Oral, lactation
1.4 pg/dL for Control
and phosphorylation
0.2% solution, M/F, n = 6-12
12.1 pg/dL for 0.2% solution
PND 30:
1.2 pg/dL for Control
12.8 pg/dL for 0.2% solution
Zhana et al. (2012)
Rat (Sprague Dawley)
NR (40-60 g)
Oral, drinking
+8 wk from start of exposure
+8 wk from start of
Control (deionized water), M,
water
exposure:
n = 10
49.9 ng/mL (5 pg/dL) for Control
Phosphorylated tau
expression, Alpha-
100 ppm, M, n = 10
100.9 ng/mL (10.1 pg/dL) for 100 ppm
Synuclein expression
200 ppm, M, n = 10
128.6 ng/mL (12.9 pg/dL) for 200 ppm
300 ppm, M, n = 10
147.7 ng/mL (14.8 pg/dL) for 300 ppm
Bihaai and Zawia (2013)
Monkey (Macaca
PND Oto PND 400
Oral, infant
PND 400:
23 yr:
fascicularis)
formula
Tau protein
Control, F, n = 4
Oral, gelatin
3-6 pg/dL for Control
expression and
capsules
phosphorylation, Tau
1.5 mg/kg/day, F, n = 5
19-26 pg/dL for 1.5 mg/kg/day
phosphorylation
3-499
-------
Study
Species (Stock/Strain), n, Timing of Exposure
Sex Exposure Details
BLL as Reported (pg/dL)
Endpoints Examined
Feng et al. (2019)
Rat (Sprague Dawley)
Control (deionized water),
M/F, n = 8
0.8 g/L (maternal) and 0.3 g/L
(pup), M/F, n = 8
1.5 g/L (maternal) and 0.9 g/L
(pup), M/F, n = 8
GD -10 to PND
490
Oral, drinking
water
Oral, lactation
In utero
PND 21:
0 mg/L (0 pg/dL) for Control
0.29 mg/L (29 pg/dL) for 0.8 g/L
0.69 mg/L (69 pg/dL) for 1.5 g/L
PND 287:
0 mg/L (0 pg/dL) for Control
0.29 mg/L (29 pg/dL) for 0.8 g/L
0.61 mg/L (61 pg/dL) for 1.5 g/L
PND 490:
0 mg/L (0 pg/dL) for Control
0.31 mg/L (31 pg/dL) for 0.8 g/L
0.58 mg/L (58 pg/dL) for 1.5 g/L
PND 21, 287, 490:
Neuronal Density,
Brain Volume
Mansouri et al. (2012)
Rat (Wistar)
Control (distilled water), M/F,
n = 16 (8/8)
50 mg/L, M/F, n = 16 (8/8)
PND 70 to PND
100
Oral, drinking
water
PND 100-Males:
2.05 pg/dL for Control
8.8 pg/dL for 50 mg/L
PND 100: Open Field
Test, Rotarod Test
PND 100 - Females:
2.17 pg/dL for Control
6.8 pg/dL for 50 mg/L
3-500
-------
Study
Species (Stock/Strain), n, Timing of Exposure
Sex Exposure Details
BLL as Reported (pg/dL)
Endpoints Examined
Mansouri et al. (2013)
Rat (Wistar)
Control (tap water or
water+NaAc), M/F, n = 16
(8/8)
50 ppm, M/F, n = 16 (8/8)
PND 55 to PND
181
Oral, drinking
water
PND 178-181 - Females:
NR for Control
10.6 pg/dL for 50 ppm
PND 155-159:
Rotarod Test
PND 178-181 - Males:
NR for Control
18.9 pg/dL for 50 ppm
Singh et al. (2019) Rat (Wistar) 3 mo to 6 mo
Control (distilled water), M, n
= 5
2.5 mg/kg, M, n = 5
Oral, gavage 6 mo:
5.76 pg/dL for Control
28.4 pg/dL for 2.5 mg/kg
6 mo: Locomotor
Activity, Rotarod Test
Al-Qahtani et al. (2022) Mouse (Albino) 8-9 wk to 14-15 wk Oral, gavage 14-15 wk: NR: Locomotor
Control (distilled water), M, n Activity
= 10 1.2 |jg/100 mL (1.2 pg/dL) for Control
0.2 mg/kg, M, n = 10
7.1 pg/100 mL (7.1 pg/dL) for 0.2
mg/kg
APP = amyloid precursor protein; BACE1 = beta-secretase 1; F = female; GD = gestational day; M = male; mo = month(s); NR = not reported; Pb = lead; PND = postnatal day; wk =
week(s); yr = year(s).
3-501
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EPA/600/R-23/375
APDA Environmental Protection Januaiy 2024
M m Agency www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 4: Cardiovascular Effects
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
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CONTENTS
DOCUMENT GUIDE 4-iii
LIST OF TABLES 4-v
LIST OF FIGURES 4-vi
ACRONYMS AND ABBREVIATION 4-viii
APPENDIX 4 CARDIOVASCULAR EFFECTS 4-1
4.1 Introduction and Summary of the 2013 Integrated Science Assessment 4-1
4.1.1 Hypertension and Increased Blood Pressure 4-2
4.1.2 Subclinical Atherosclerosis 4-4
4.1.3 Coronary Heart Disease 4-5
4.1.4 Cerebrovascular Disease 4-6
4.2 Scope 4-7
4.3 Blood Pressure and Hypertension 4-8
4.3.1 Epidemiologic Studies of Blood Pressure and Hypertension 4-8
4.3.2 Toxicological Studies of Blood Pressure and Hypertension 4-36
4.3.3 Integrated Summary of Blood Pressure and Hypertension 4-38
4.4 Ischemic Heart Disease and Associated Cardiovascular Effects 4-40
4.4.1 Epidemiologic Studies of Ischemic Heart Disease 4-40
4.4.2 Summary of Ischemic Heart Disease 4-44
4.5 Heart Failure and Impaired Cardiac Function 4-45
4.5.1 Epidemiologic Studies of Impaired Cardiac Function 4-45
4.5.2 Toxicological Studies of Impaired Cardiac Function 4-46
4.5.3 Integrated Summary of Impaired Cardiac Function 4-47
4.6 Endothelial Dysfunction 4-48
4.6.1 Toxicological Studies of Endothelial Dysfunction 4-48
4.6.2 Summary of Endothelial Dysfunction 4-49
4.7 Cardiac Electrophysiology and Arrythmia 4-49
4.7.1 Cardiac Depolarization, Repolarization, and Arrythmia 4-49
4.7.2 Heart Rate and Heart Rate Variability 4-51
4.7.3 Integrated Summary of Cardiac Electrophysiology and Arrythmia 4-53
4.8 Atherosclerosis and Peripheral Artery Disease 4-54
4.8.1 Epidemiologic Studies of Atherosclerosis and Peripheral Artery Disease 4-54
4.8.2 Toxicological Studies of Atherosclerosis 4-57
4.8.3 Integrated Summary of Atherosclerosis 4-58
4.9 Cerebrovascular Disease 4-58
4.9.1 Epidemiologic Studies of Cerebrovascular Disease 4-58
4.9.2 Summary of Cerebrovascular Disease 4-59
4.10 Cardiovascular Mortality 4-59
4.10.1 Epidemiologic Studies of Cardiovascular Mortality 4-59
4.10.2 Summary of Cardiovascular Mortality 4-67
4.11 Biological Plausibility 4-68
4.12 Summary and Causality Determination 4-72
4.13 Evidence Inventories - Data Tables to Summarize Study Details 4-80
4.14 References 4-151
4-iv
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LIST OF TABLES
Table
4-1
Summary of causality determinations from the 2013 Pb Integrated Science Assessment
4-2
Table
4-2
Summary of evidence indicating a causal relationship between Pb exposure and
cardiovascular effects and cardiovascular-related mortality
4-77
Table
4-3
Epidemioloqic studies of Pb exposure and blood pressure
4-80
Table
4-4
Epidemioloqic studies of Pb exposure and hypertension
4-97
Table
4-5
Epidemiologic studies of Pb exposure and blood pressure and hypertension among
children
4-107
Table
4-6
Animal toxicoloqical studies of Pb exposure and blood pressure/hypertension
4-114
Table
4-7
Epidemioloqic studies of Pb exposure and coronary and ischemic heart disease
4-118
Table
4-8
Epidemioloqic studies of Pb exposure and cardiac function
4-125
Table
4-9
Animal toxicoloqical studies of cardiac function
4-128
Table
4-10
Animal toxicoloqical studies of Pb exposure and endothelial dysfunction
4-130
Table
4-11
Epidemioloqic studies of Pb exposure cardio electrophysioloqy and arrythmia
4-131
Table
4-12
Animal toxicoloqical studies of Pb exposure and cardiac electrophysioloqy
4-135
Table
4-13
Epidemiologic studies of Pb exposure and atherosclerosis and peripheral artery disease
4-137
Table
4-14
Animal toxicoloqical studies of Pb exposure and atherosclerosis
4-140
Table
4-15
Epidemioloqic studies of Pb exposure and cerebrovascular disease
4-141
Table
4-16
Epidemioloqic studies of Pb exposure and cardiovascular mortality
4-143
4-v
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LIST OF FIGURES
Figure 4-1 Association between biomarkers of Pb exposure and blood pressure. 4-12
Figure 4-2 Association between blood Pb level quartiles and systolic blood pressure, diastolic blood
pressure, and hypertension, polluted region of rural southwest China. 4-13
Figure 4-3 Association between blood Pb level quartiles and systolic blood pressure, diastolic blood
pressure, and hypertension, unpolluted region of rural southwest China. 4-14
Figure 4-4 Effect measure modification by sex and race for blood pressure (systolic and diastolic)
and a doubling of blood Pb levels, National Health and Nutrition Examination Survey
(2003-2010). 4-18
Figure 4-5 Effect measure modification by sex and race for blood pressure (systolic, diastolic, and
pulse pressure) and a doubling of blood Pb level, National Health and Nutrition
Examination Survey (1999-2006). 4-19
Figure 4-6 Effect measure modification by sex and race for blood pressure (systolic, diastolic, and
pulse pressure) and a doubling of blood Pb levels, National Health and Nutrition
Examination Survey (2001-2008). 4-20
Figure 4-7 Effect measure modification between blood Pb levels, race, and education level, National
Health and Nutrition Examination Survey (2001-2008). 4-21
Figure 4-8 Effect measure modification between blood Pb levels, race, and poverty level, National
Health and Nutrition Examination Survey (2001-2008). 4-22
Figure 4-9 Effect measure modification by sex and age of the relationship between blood Pb levels
and systolic blood pressure, Canadian Health Measures Survey. 4-24
Figure 4-10 Effect measure modification by sex and age of the relationship between blood Pb levels
and diastolic blood pressure, Canadian Health Measures Survey. 4-25
Figure 4-11 Effect measure modification by vitamin D receptor variant for the association between
pulse pressure and tibial Pb levels, Normative Aging Study cohort. 4-26
Figure 4-12 Associations between biomarkers of Pb exposure and hypertension. 4-28
Figure 4-13 Dose-response curve between tibia Pb levels and resistant hypertension, Normative Aging
Study cohort. 4-31
Figure 4-14 Dose-response curve between blood Pb and any hypertension or uncontrolled
hypertension, restricted cubic splines, National Health and Nutrition Examination Survey
(1999-2006). 4-32
Figure 4-15 Effect measure modification by sex for the association between quartiles of blood Pb and
prevalent hypertension. 4-33
Figure 4-16 Relationship between blood Pb levels and common carotid artery plaques, common
carotid artery diameter, and cardiovascular disease among diabetic patients. 4-43
Figure 4-17 Meta-analysis of the association between biomarkers of Pb exposure and coronary heart
disease. 4-44
Figure 4-18 Association between aortic pulse wave velocity with blood Pb levels and age. 4-56
4-vi
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Figure 4-19 Stratified associations between abdominal aortic calcification score and blood Pb levels. 4-57
Figure 4-20 Associations between blood Pb level and cardiovascular mortality. 4-60
Figure 4-21 Dose-response relationship between blood Pb levels and cardiovascular and ischemic
heart disease mortality. 4-64
Figure 4-22 Cumulative incidence function of cardiovascular mortality by blood Pb level, National
Health and Nutrition Examination Survey III (1988-1994). 4-65
Figure 4-23 Potential biological pathways for cardiovascular effects following exposure to Pb. 4-69
4-vii
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ACRONYMS AND ABBREVIATION
A peak late diastolic velocity
AAC abdominal aortic calcification
AAS atomic absorption spectrometry
ABLES Adult Blood Lead Epidemiology and
Surveillance
ACE angiotensin-converting enzyme
ACh acetylcholine
ADMA asymmetric dimethylarginine
AF atrial fibrillation
AGT angiotensinogen
AHA American Heart Association
AL allostatic load
ALAD S-aminolevulinic acid dehydratase
ANS autonomic nervous system
APOE apolipoprotein E
AQCD Air Quality Criteria Document
ARCA Automobile Racing Club of America
ASCVD atherosclerotic cardiovascular disease
ASV anodic stripping voltammetry
BEST Bangladesh Vitamin E and Selenium
Trial
BLL blood lead level
BMI body mass index
BP blood pressure
BRHS British Regional Heart Study
BW body weight
Ca2+ calcium ion(s)
C282Y HFE mutant of the HFE wildtype
CAD coronary artery disease
CAS coronary artery stenosis
CCA common carotid artery
Cd cadmium
CHD coronary heart disease
CHF congestive heart failure
CI confidence interval
CIF cumulative incidence function
CRP C-reactive protein
CT computerized tomography
CVD cardiovascular disease
d day(s)
DBP diastolic blood pressure
E peak early diastolic velocity
e' peak early diastolic mitral annular
velocity
EAF electric arc furnace
ECG electrocardiography
EDTA ethylenediaminetetraacetic acid
eGFR estimated glomerular filtration rate
EMM effect measure modification
ETAAS Electrothermal Atomic Absorption
Spectrometry
Eyr erythrocyte
FABP4 adipocyte fatty acid-binding protein
FBG fasting blood glucose
Fe iron
FRS Framingham risk score
GFAAS graphite furnace atomic absorption
spectrometry
GFR glomerular filtration rate
GLS global longitudinal strain
GM geometric mean
GRS genetic risk score
GSD geometric standard deviation
GSE geometric standard error
GuLF Gulf Long-T erm F ollow-up
GW gestational week
FlbAlc hemoglobin Ale
H63D HFE mutant of the HFE
HDL high-density lipoprotein
HDL-C high-density lipoprotein cholesterol
HF high frequency
HFE hemochromatosis gene
HMOX1 heme oxygenase-1
HOME Health Outcomes and Measures of the
Environment
hr hour(s)
HR hazard ratio
HRV heart rate variability
HTN hypertension
ICD International Classification of Diseases
ICP-MS inductively coupled plasma mass
spectrometry
IHD ischemic heart disease
IMT intimal medial thickening
IQR interquartile range
ISA Integrated Science Assessment
IVS interventricular septum
KNHANES Korea National Health and Nutrition
Examination Survey
K-XRF K-shell X-ray fluorescence
LCL lower confidence limit
4-viii
-------
LDL low-density lipoprotein
LDL-C low-density lipoprotein cholesterol
LF low frequency
LV left ventricular
LVDP left ventricular diastolic pressure
LVMI left ventricular mass index
LVPW left ventricular posterior wall
LVSP left ventricular systolic pressure
MAP mean arterial pressure
MDCS-CC cardiovascular cohort of the Malmo
Cancer and Diet Study
METAL Environmental Pollutant Exposure and
Metabolic Diseases in Shanghai
METS Modeling the Epidemiologic Transition
Study
MI myocardial infarction
mo month(s)
NA not available
NAS Normative Aging Study
NASCAR National Association for Stock Car
Auto Racing
NH non-Hispanic
NHANES National Health and Nutrition
Examination Survey
NN normal-to-normal
NO nitric oxygen
NR not reported
nu normalized units
OR odds ratio
PAD peripheral artery disease
Pb lead
PECOS Population, Exposure, Comparison,
Outcome, and Study Design
PHQ Patient Health Questionnaire
PIR poverty-income ratio
PIVUS Prospective Investigation of the
Vasculature in Uppsala Seniors
PND postnatal day
PP pulse pressure
PR prevalence ratio
PROGRESS Programming Research in Obesity,
Growth Environment and Social
Stressors
PVD peripheral vascular disease
PWV pulse wave velocity
Q quartile
QRS QRS complex in ECG
QRSc corrected QRS duration
QT QT interval in ECG
QTc corrected QT interval
RAAS renin-angiotensin-aldosterone system
RCT randomized control trial
RLS regional longitudinal strain
rMSSD root-mean-square of successive
differences
ROS reactive oxygen species
RR relative risk
RRS regional radial strain
RV right ventricular
RVDP right ventricular diastolic pressure
RVSP right ventricular systolic pressure
RWT relative wall thickness
SBP systolic blood pressure
SD standard deviation
SE standard error
SES socioeconomic status
SNP single nucleotide polymorphism
SOD superoxide dismutase
SOF Study of Osteoporotic Fractures
ST segment measured from the J point to
the end of the T-wave in an ECG
T# tertile #
TACT Trial to Assess Chelation Therapy
TC total cholesterol
TPR total peripheral resistance
TRI Toxic Release Inventory
UCL upper confidence limit
VA-NAS Veterans Affairs Normative Aging
Study
VDR vitamin D receptor
yr year(s)
4-viii
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APPENDIX 4 CARDIOVASCULAR EFFECTS
Causality Determination for Lead (Pb) Exposure and
Cardiovascular Effects and Cardiovascular-Related Mortality
This appendix characterizes the scientific evidence that supports causality determinations for
lead (Pb) exposure and cardiovascular effects. The types of studies evaluated within this appendix are
consistent with the overall scope of the ISA as detailed in the Process Appendix (see Section 12.4). In
assessing the overall evidence, the strengths and limitations of individual studies were evaluated based
on scientific considerations detailed in Table 12-5 of the Process Appendix (see Section 12.6.1). More
details on the causal framework used to reach these conclusions are included in the Preamble to the
ISA (U.S. EPA, 2015). The evidence presented throughout this appendix supports the following
causality conclusion:
Outcome
Causality Determination
Cardiovascular Effects and
Causal
Cardiovascular-Related Mortality
The Executive Summary, Integrated Synthesis, and all other appendices of this Pb ISA can be found at
https://asscssmcnts.cpa.go\/isa/documcnt/&dcid=359536.
4.1 Introduction and Summary of the 2013 Integrated Science
Assessment
The 2013 Integrated Science Assessment for Lead (hereinafter referred to as the 2013 Pb ISA)
(U.S. EPA, 2013) made four causality determinations with respect to cardiovascular disease, using the
U.S. Surgeon General's Report on Smoking as a guideline to group evidence into health outcome
categories (CDC, 2004). The categories included hypertension, subclinical atherosclerosis, coronary heart
disease (CHD), and cerebrovascular disease. Evidence was sufficient to conclude causal relationships
between Pb exposure and hypertension and CHD. The causal determination for hypertension was not only
informed by evidence of hypertension and increases in blood pressure (BP), but also cardiovascular-
related mortality. The 2013 Pb ISA indicated a coherence between epidemiologic and toxicological
studies, and animal toxicological studies provided strong evidence to support biologic plausibility.
Specifically, oxidative stress from Pb exposure can result in an inactivation of nitrous oxide, which can
lead to increased vasoconstriction and therefore increased BP. The causal determination for CHD was
informed by epidemiologic evidence for heart rate variability (HRV); myocardial infarction (MI);
ischemic heart disease (IHD); mortality from MI, IHD, and CHD; and increased thrombosis, coagulation,
and arrhythmia in animals. The biological plausibility and mode of action for these cardiovascular effects
was provided by evidence for oxidative stress, inflammation, and vascular cell activation or dysfunction.
Specifically, coherence between epidemiologic and toxicologic evidence demonstrated that Pb exposure
4-1
-------
may promote a procoagulant state that can potentially contribute to thrombus formulation and therefore
reduced blood supply to the heart. Causality determinations for each of the four categories are
summarized in Table 4-1 and some of the evidence supporting these determinations is discussed in
Sections 4.1.1 to 4.1.4.
Table 4-1 Summary of causality determinations from the 2013 Pb Integrated
Science Assessment
Outcome Group
Causality Determination
Hypertension and Increased Blood Pressure
Causal
Subclinical Atherosclerosis
Suggestive
Coronary Heart Disease
Causal
Cerebrovascular Disease
Inadequate
The current ISA is consistent with more recent ISAs (e.g., 2019 Particulate Matter and 2020
Ozone ISAs) (U.S. EPA. 2020. 2019) in that it makes a single causality determination for cardiovascular
effects. This approach recognizes that many cardiovascular endpoints are inter-related (e.g., both
atherosclerosis and endothelial dysfunction can contribute to increases in BP), and therefore not easily
discussed in isolation. The remainder of this section summarizes the evidence for Pb exposures and
cardiovascular effects assessed in the 2013 Pb ISA, including the evidence for hypertension and increased
BP (Section 4.1.1), atherosclerosis (Section 4.1.2), CHD (Section 4.1.3), and cerebrovascular disease
(Section 4.1.4). Subsequent sections of this appendix provide an overview of study inclusion criteria for
the cardiovascular effects evidence in the current ISA (Section 4.2), summaries and evaluations of recent
health effects evidence (Sections 4.3 to 4.10), a discussion of biological plausibility (Section 4.1.1), and a
discussion of how all the individual lines of cardiovascular evidence were considered and integrated to
inform the causality determination for Pb exposure and cardiovascular effects (Section 4.1.2). Study-
specific details, including information on study design; exposure metrics, concentrations, and durations;
and select results are presented in summary tables in Section 4.1.3.
4.1.1 Hypertension and Increased Blood Pressure
The 2013 Pb ISA (U.S. EPA, 2013) indicated that the combined evidence from epidemiologic
and animal toxicological studies was sufficient to conclude that there is a causal relationship between Pb
exposure and hypertension. This conclusion was informed by the coherence of effects observed between
epidemiologic and toxicological studies with respect to hypertension and its related endpoints. A number
of prospective epidemiologic studies clearly supported the relationship between biomarkers of Pb
exposure and hypertension incidence and changes in BP (U.S. EPA. 2013). The prospective evidence
4-2
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was supported by meta-analyses that underscored the consistency and reproducibility of Pb-associated
increases in BP and hypertension across diverse populations and different study designs. Importantly,
many epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA, 2013) adjusted for a wide range of
potential confounders to reduce uncertainty due to potential unmeasured confounding. Although the
adjustment for specific factors varied by study, the collective body of evidence included adjustments for
multiple potential key confounding factors, including age, diet, sex, body mass index (BMI), BP-lowering
medication use, socioeconomic status (SES), race/ethnicity, alcohol consumption, cholesterol, smoking,
preexisting disease (e.g., diabetes), measures of renal function, and copollutant exposures (e.g., cadmium
[Cd]), while still maintaining positive associations between biomarkers of Pb exposure and changes in
BP/hypertension.
Results from animal toxicological studies examining BP-related endpoints were consistent with
the epidemiologic findings. In the previous review, all the animal toxicological studies providing blood
Pb level (BLL) and BP measurements reported increases in BP with increasing BLLs. Most of these
studies examined Pb exposures that resulted in mean BLLs >10 (ig/dL; however, a single animal
toxicological study conducted in rats after drinking water exposure found a continuous increase in BP in
animals with mean BLLs ranging from 0.05 to 29 (ig/dL with no evidence of a threshold (U.S. EPA,
2013).
Animal toxicological studies also provided strong support for the biological plausibility of the Pb-
associated increases in BP and hypertension observed in epidemiologic studies. Studies evaluated in the
2013 Pb ISA (U.S. EPA, 2013) demonstrated that oxidative stress following Pb exposure inactivates
vasodilator nitric oxide, which may lead to increased vasoconstriction and increased BP. If increases in
BP persist, the result is hypertension (i.e., chronically elevated BP). Oxidative stress can also damage the
endothelium, further disrupting endothelium-dependent vascular relaxation and increasing the contractile
response. Studies also suggested Pb exposure disrupts normal contractile processes by altering the
sympathetic nervous system, the renin-angiotensin-aldosterone system, and the balance between
production of vasodilators and vasoconstrictors (U.S. EPA, 2013).
Although the relationship between exposure to Pb and increases in BP in adults was well
established at the time of the last review, some uncertainties were identified in the evidence for BP
changes, specifically among children. The 2013 Pb ISA noted that some of the BP results (and other
cardiovascular effects) observed in children may be antecedent to later-in-life effects. Therefore, there is
at least some uncertainty in the level, timing, frequency, and duration of Pb exposure contributing to the
reported cardiovascular effects in adults. That is, although there is a clear relationship between exposure
to Pb and changes in BP in adults, it is possible that childhood Pb exposures could contribute to adult
BLLs through processes such as bone remodeling that occurs during aging and/or pregnancy. Thus, Pb-
associated changes in BP reported in adults may be appreciably influenced by past Pb exposures, perhaps
as early as childhood.
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Overall, epidemiologic and toxicological evidence from the previous review consistently
demonstrated that Pb exposures are associated with increased BP and hypertension in adults. The
epidemiologic studies have been replicated by different researchers in different cohorts and associations
reported in these studies have largely remained positive after adjusting for numerous potential
confounding factors. These studies are also coherent with numerous animal toxicological studies
demonstrating increases in BP following Pb exposure. In addition, toxicological studies provided
biological plausibility and a potential mode of action for the results observed in epidemiologic studies.
Thus, in the 2013 Pb ISA, the combined evidence from epidemiologic and animal toxicological studies
was sufficient to conclude that there is a causal relationship between Pb exposure and hypertension.
Since the last review, the evidence relating Pb exposure to increases in BP and hypertension have
expanded greatly (see Section 4.3), further reinforcing an already strong evidence base established in the
last ISA. As a result, evidence from epidemiologic and animal toxicological studies related to BP and
hypertension are a key driver for the current ISA's conclusion of a causal relationship between Pb
exposure and cardiovascular effects. A discussion of how the epidemiologic and animal toxicological
evidence of hypertension and increased BP contributes to the determination of a causal relationship
between exposure to Pb and cardiovascular effects in this ISA can be found in Section 4.1.2.
4.1.2 Subclinical Atherosclerosis
The 2013 Pb ISA (U.S. EPA, 2013) concluded that the evidence was suggestive of, but not
sufficient to infer, a causal relationship between exposure to Pb and subclinical atherosclerosis. Studies
considered in the last review included an analysis from the 2006 Pb Air Quality Criteria Document
(AQCD) indicating that exposure to Pb was associated with peripheral artery disease (PAD) in the
National Health and Nutrition Examination Survey (NHANES) population, and that co-exposure with Cd
did not confound the association (U.S. EPA, 2006). Additional epidemiologic findings presented in the
2013 Pb ISA were limited to cross-sectional analyses. One such analysis reported an increasing trend in
the odds of PAD across concurrent BLL groups in adult NHANES participants. Furthermore, in an
epidemiologic study conducted in a Pb-exposed population, positive associations were reported between
BLLs and increases in intima-media thickness and atherosclerotic plaque presentation (U.S. EPA, 2013).
However, the 2013 Pb ISA noted that most of the available epidemiologic analyses were cross-sectional
in nature, contributing to uncertainty in the level, timing, frequency, and duration of the Pb exposures that
contributed to the observed associations.
In addition to the epidemiologic evidence, toxicological studies provided limited evidence to
suggest Pb exposure may initiate atherosclerotic vessel disease. The 2013 Pb ISA (U.S. EPA, 2013)
noted that in vitro Pb exposure resulted in a concentration-dependent increase in arterial intimal thickness
in human radial and internal mammary arteries. Moreover, exposure to Pb in rats increased aortic medial
thickness. Following Pb exposure, toxicological studies also demonstrated evidence of oxidative stress
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and systemic inflammation, processes which are important to the development of atherosclerosis. Finally,
toxicological studies indicated a relationship between Pb exposure and elevation of cholesterol (U.S.
EPA, 2013). Taken together with the epidemiologic evidence and its associated uncertainties, the 2013
Pb ISA concluded that the evidence was suggestive of a causal relationship between Pb exposure and
subclinical atherosclerosis.
Since the last review, there is additional evidence supporting the potential contribution of Pb
exposures to atherosclerosis. This evidence includes recent epidemiologic studies reporting positive
associations with markers of atherosclerosis and a recent toxicological study in rats demonstrating
morphological changes in the aorta consistent with the potential for atherosclerosis (see Section 4.3). A
discussion of how the epidemiologic and animal toxicological evidence of atherosclerosis contributes to
the determination of a causal relationship between exposure to Pb and cardiovascular effects in this ISA
can be found in Section 4.1.2.
4.1.3 Coronary Heart Disease
The 2013 Pb ISA (U.S. EPA, 2013) concluded that the evidence supports a causal relationship
between exposure to Pb and CHD. This conclusion was primarily based on the results of epidemiologic
studies examining the incidence of MI, IHD, and HRV, and on studies examining mortality from CHD,
MI, or IHD. The rationale for this determination is summarized below.
The 2013 Pb ISA (U.S. EPA, 2013) described longitudinal studies in cohorts in different
locations with follow-up periods of up to 12 years. These studies consistently reported that biomarkers of
Pb exposure are associated with risk of mortality from MI, IHD, or CHD. The strongest associations were
observed with MI mortality. Despite the differences in design and methods used across epidemiologic
studies, associations between higher levels of tissue Pb (blood, bone) and higher risk of CHD-related
mortality were generally observed 2013 Pb ISA (U.S. EPA, 2013). The body of evidence demonstrating
associations with mortality from CHD was substantiated by several studies indicating associations
between biomarkers of Pb exposure and incidence of CHD-related outcomes. For example, a prospective
analysis examined the incidence of IHD (physician-confirmed MI, angina pectoris) and reported that
blood and bone Pb levels contributed independently to IHD incidence. The 2013 Pb ISA further noted
that coherence for the associations in humans was provided by an animal toxicological study suggesting
that Pb exposure promoted a procoagulant state that could contribute to thrombus formation, and thus,
potentially reduce the blood supply to the heart (U.S. EPA, 2013).
Previous research has indicated that decreased HRV is associated with higher mortality from MI
and can be used as a predictor of the physiological processes underlying CHD. The 2013 Pb ISA
described several cohort studies demonstrating associations between Pb exposure and decreases in HRV
(U.S. EPA, 2013). Additionally, a prospective analysis reported that higher tibia Pb levels were
associated with increases in certain ECG measurements, including the corrected QT interval (QTc) and
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corrected QRS duration (QRSc), which can be indicative of impaired cardiac electrophysiology. As CHD
is the result of vascular blockage, the previous Pb ISA also noted that these epidemiologic associations
were supported, at least in part, by the limited evidence for subclinical atherosclerosis (Section 4.1.2).
The 2013 Pb ISA (U.S. EPA, 2013) additionally noted that the strong and consistent evidence for Pb-
induced hypertension supported the biological plausibility of the Pb-induced increase in CHD risk.
In summary, in the 2013 Pb ISA (U.S. EPA, 2013), several studies examining CHD morbidity
and mortality and contributing cardiovascular effects reported consistent associations between Pb
exposure and CHD. In addition, both animal toxicological and epidemiologic studies describe a
biologically plausible potential mode of action (e.g., hypertension, atherosclerosis, potentially adverse
changes in cardiac electrophysiology). Taken together, the 2013 Pb ISA (U.S. EPA, 2013) concluded that
epidemiologic evidence, supported by toxicological evidence, was sufficient to conclude a causal
relationship exists between Pb exposure and CHD.
Since the last review, the epidemiologic evidence describing the relationship between Pb
exposure and endpoints related to CHD has expanded (see Section 4.4) and further strengthens the
evidence base established in the last ISA. In particular, there are new epidemiologic studies reporting
associations with IHD and MI mortality. Results of animal toxicological studies of HRV and cardiac
electrophysiology published since the last review have been largely mixed. A discussion of how this
epidemiologic and animal toxicological evidence contributed to the determination of a causal relationship
between exposure to Pb and cardiovascular effects in this current review can be found in Section 4.1.2.
4.1.4 Cerebrovascular Disease
The 2013 Pb ISA (U.S. EPA, 2013) concluded there was insufficient evidence to inform the
relationship between cerebrovascular disease and Pb exposure. Despite strong evidence indicating effects
of Pb exposure on hypertension and CHD, very few studies evaluated in the 2013 Pb ISA examined the
effects of Pb exposure on cerebrovascular disease. Furthermore, the studies that were available reported
relatively imprecise associations between BLLs and stroke-related mortality. With respect to animal
toxicological studies, there was some evidence for processes that could lead to cerebrovascular disease,
such as an increase in markers of oxidative stress, inflammation, and coagulation that could potentially
aid in clot formation. When considered as a whole, however, this limited evidence was insufficient to
inform the relationship between Pb exposure and cerebrovascular disease. In the current review, studies
examining the potential relationship between Pb exposure and cerebrovascular disease remain quite
limited (see Section 4.9). Consideration of this evidence in the causality determination for Pb exposures
and cardiovascular effects is presented in Section 4.1.2.
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4.2 Scope
The scope of this appendix is defined by Population, Exposure, Comparison, Outcome, and Study
Design (PECOS) statements. The PECOS statement defines the objectives of the review and establishes
study inclusion criteria and thereby facilitates identification of the most relevant literature to inform the
Pb ISA.1 To identify the most relevant literature, the body of evidence from the 2013 Pb ISA was
considered in the development of the PECOS statements for this appendix. Specifically, well-established
areas of research; gaps in the literature; and inherent uncertainties in specific populations, exposure
metrics, comparison groups, and study designs identified in the 2013 Pb ISA inform the scope of this
appendix. The 2013 Pb ISA used different inclusion criteria than the current ISA, and many of the studies
referenced therein do not meet the current PECOS criteria (e.g., due to higher or unreported biomarker
levels). Many of those studies are discussed in this appendix to establish the state of the evidence prior to
this assessment. With the exception of supporting evidence used to demonstrate the biological plausibility
of Pb-associated cardiovascular effects, recent studies were only included if they satisfied all of
components of the following discipline-specific PECOS statements:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect;
Exposure: Exposure to Pb2 as indicated by biological measurements of Pb in the body—with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure,3 or intervention groups in randomized trials and quasi-experimental studies;
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles);
Outcome: Cardiovascular effects including but not limited to CHD, hypertension and increased
BP, and cardiovascular-related mortality; and
'The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
2Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area that was of
particular relevance to the National Ambient Air Quality Standards review (e.g., longitudinal studies designed to
examine recent versus historical Pb exposure).
3Studies that estimate Pb exposure by measuring Pb concentrations in PMw and PM2.5 ambient air samples are only
considered for inclusion if they also include a relevant biomarker of exposure. Given that size distribution data for
Pb-PM are fairly limited, it is difficult to assess the representativeness of these concentrations to population
exposure [Section 2.5.3 (U.S. EPA. 2013)1. Moreover, data illustrating the relationships of Pb-PMio and Pb-PMis
with BLLs are lacking.
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Study design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages);
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a BLL of 30 (ig/dL or below;4'5
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated control
Outcome: Cardiovascular effects; and
Study design: Controlled exposure studies of animals in vivo.
4.3 Blood Pressure and Hypertension
High BP typically is defined as a systolic BP (SBP) above 130 mmHg or a diastolic blood
pressure (DBP) above 80 mmHg. SBP represents the pressure in the arteries as the heart contracts, while
DBP represents the pressure in the arteries as the heart is relaxed and is filling with blood. Prolonged high
BP is known as hypertension and can lead to a thickening of the ventricular wall resulting in diminished
filling during diastole. Hypertension can contribute ultimately to the development of arrythmia and heart
failure. Pulse pressure (PP), or the difference between SBP and DBP, as well as mean arterial pressure
(MAP)—which is a function of cardiac output, systemic vascular resistance, and central venous
pressure—are additional metrics used in studies of air pollution's effects on BP. Moreover, hypertension
is one of several conditions, including high blood sugar, excess body fat around the waist, and abnormal
triglyceride levels, that comprise metabolic syndrome (see Appendix 9), which is a risk factor for heart
disease, stroke, and diabetes.
4.3.1 Epidemiologic Studies of Blood Pressure and Hypertension
Several epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA, 2013) and previous
AQCD documents (U.S. EPA, 2006, 1990) indicate an association between biomarkers of Pb exposure
4Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
5This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLL.
The 95th percentile of the 2011-2016 NHANES distribution of BLL in children (1-5 years; n= 2,321) is 2.66 (ig/dL
(Egan et al.. 2021) and the proportion of individuals with BLL that exceed this concentration varies depending on
factors including (but not limited to) housing age, geographic region, and a child's age, sex, and nutritional status.
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and changes in BP and hypertension risk. Although previous studies evaluated in the 2006 Pb AQCD
(U.S. EPA, 2006) and a supplement to the 1986 Pb AQCD (U.S. EPA, 1990) most likely represented
populations historically exposed to higher levels of air Pb (measured during the 1970s and 1980s)
compared with populations today, they indicated there was no apparent threshold below which blood Pb
was not significantly associated with changes in BP, for mean BLLs ranging from 7 (ig/dL to 34 (ig/dL.
The 2013 Pb ISA (U.S. EPA, 2013) further demonstrated an association between Pb biomarkers and
increased BP and hypertension risk at BLLs <2 (ig/dL. The majority of the evidence for this association
was derived from the Normative Aging Study (NAS) cohort of mostly older white men (Zhang et al.,
2010; Perlstein et al., 2007; Elmarsafawy et al„ 2006) and a Korean study composed of workers with high
BLLs (mean BLLs -20-35 (.ig/dL). due to occupational Pb exposure (Weaver et al., 2008; Glenn et al.,
2006). The 2013 Pb ISA (U.S. EPA, 2013) also highlighted specific groups that may be at higher risk of
an adverse BP outcome with increased Pb biomarkers, including those with high stress, certain genetic
variants, and minority populations.
Recent studies continue to provide consistent evidence that exposure to Pb is associated with
increased BP and hypertension risk. The majority of recent studies evaluating Pb biomarkers and changes
in BP or hypertension are cross-sectional, which can be useful for assessing concurrent associations
between blood Pb and increased BP or hypertension risk. A smaller number of studies implemented a
longitudinal study design, useful for evaluating long-term effects of elevated Pb biomarkers. Generally,
the evidence continues to indicate that changes in BP are most strongly associated with concurrent BLLs,
whereas increased risk of hypertension is more likely to be associated with cumulative Pb measures (such
as bone Pb levels).
4.3.1.1 Blood Pressure
Several recent studies specifically evaluated SBP and DBP, while other studies examined changes
in PP and MAP. Study-specific details, including blood/bone Pb levels, study population characteristics,
confounders, and selected results from these studies, are highlighted in Table 4-3. Studies in Figure 4-1
are standardized to be interpreted as changes in BP associated with a 1 (ig/dL increase in BLL or a
10 (ig/g increase in bone Pb level. Study details in Table 4-3 include standardized results as well as results
that could not be standardized on the basis of information provided in each paper. Many of these studies
evaluated this association cross-sectionally. Specifically, most used population-level cross-sectional study
designs using NHANES (Huang. 2022; Everson et al.. 2021; Teve et al.. 2020; Obcng-Gvasi. 2019;
Obeng-Gvasi et al.. 2018; Hara et al.. 2015; Hicken et al.. 2013; Zota et al.. 2013a; Hicken et al.. 2012;
Scinicariello et al.. 2011). Korea National Health and Nutrition Examination Survey (KNHANES) (Lee et
al.. 2016a). a Canadian population-level survey (Canadian Health Measures Survey) (Bushnik et al..
2014). or a Chinese longitudinal survey (China National Human Biomonitoring) (Qu et al.. 2022). These
types of population-level cross-sectional studies have the advantage of assessing relatively low average
blood Pb (<5 (ig/dL) levels with concurrent BP measurements among a large sample size of participants.
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A single study (Scinicaricllo et al.. 2010) used this type of data in the 2013 Pb ISA to evaluate BLLs and
changes in BP measurements. Additionally, two recent studies longitudinally evaluated the association
between biomarkers of Pb exposure and BP changes (Yu et al.. 2020; Bulka et al.. 2019).
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Reference Population
SBP:
tAlmeida Lopes et al, 2017 Adults >40 Cambe. Brazil
tHuarig et al, 2022
tEverson et al, 2021
Glenn et al, 2006
Weaver et al, 2008
Scinicariello et al, 2010
NHANES
All
Men
Mexican American
Other Hispanic
Non-Hispanic White
Non-Hispanic Black
Other Race
Women
Mexican American
Other Hispanic
Non-Hispanic White
Non-Hispanic Black
Other Race
NHANES
NH White
NH Black
Hispanic
Other
NHANES
Pb distribution Pb biomarker
Geometric mean: 1.97 (95%CI:1.90-2.04) Blood
Mean (SD): 1.73 (1.71) Blood
Median (IQR)
Men: 1.50 (0.99, 2.29)
Women-1 06 (0 69,1 60)
Men: 1 60 (1.00,2 60)
Women: 1.11 (0.71,1.77)
Men: 1.58(0.99, 2.43)
Women 0 95 (0.62,1.51)
Men: 1.54(1.05.2.39)
Women. 1.16(0.75, 1.79)
Median 1.5
Korean Pb Workers
Short term: Longitudinal blood Pb
Short term: Concurrent blood Pb
Long term: Longitudinal blood Pb
Long term: Concurrent blood Pb
Korean Pb Workers
Blood
Blood
Mean (SD): 30.0(16.7)
Mean (SD): 75.1 (101.1)
Mean (SE)
Overall: 2.99 (0.99)
Non-Hispanic White: 2.87 (0.09)
Non-Hispanic Black 3.59 (0.20)
Mexican American 3.33 (0.11)
tAlmeida Lopes et al, 2017 Adults, >40 Cambe, Brazil
tHuang et al, 2022 NHANES
All
Men
Mexican Amencan
Other Hispanic
Non-Hispanic White
Non-Hispanic Black
Other Race
Women
Mexican American
Other Hispanic
NorvHispanic White
Non-Hispanic Black
Other Race
tEverson et al, 2021
Zhang etal, 2010
Geometric mean: 1.97 (95%CI: 1 90-2.04) Blood
Mean (SD): 1.73 (1.71) Blood
NHANES
Median (IQR)
Blood
NH White
Men: 1.50(0.99,2.29)
Women 1 06 (0.69. 1.60)
NH Black
Men: 1.60 (1.00, Z60)
Women 1.11 (0.71,1.77)
Hispanic
Men 1 58 (0.99,2.43)
Women: 0.95 (0.62,1.51)
Other
Men: 1.54(1.05,2.39)
Women 1 16(0 75,1.79)
NHANES
Median: 1.5
Blood
NHANES III
Mean (SE)
Blood
Overall: 2 99 (0.99)
Non-Hispanic White: 2.87 (0.09)
Non-Hispanic Black 3.59 (0.20)
Mexican American 3.33 (0.11)
NAS men
HFE Wildtype
18(12-27)
Tibia
HFE H63D
19(14-26)
Tibia
HFE C282Y
20(14-27)
Tibia
Any HFE variant
19(14-27)
Tibia
HFE Wildtype
26(17-34)
Patella
HFE H63D
27(19-37)
Patella
HFE C282Y
25(17-37)
Patella
Any HFE variant
26(18-37)
Patella
0.00 0.50 1.00 1.50 2.00
in BP (mmHg 95% CI) per 1 ug/dL increase in blood Pb or 10 ug/dL increase in bone Pb
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Figure 4-1 (Continued) Association between biomarkers of Pb exposure and
blood pressure.
AL = allostatic load; BP = blood pressure; HFE C282Y = mutant of the HFE wildtype; CI = confidence interval; DBP = diastolic blood
pressure; GSE = geometric standard error; HFE H63D = mutant of the HFE wildtype; HFE = hemochromatosis gene;
IQR = interquartile range; NAS = Normative Aging Study; NH = non-Hispanic; NHANES = National Health and Nutrition Examination
Survey; OR = odds ratio; Pb = lead; Q# = quartile number; RR = relative risk; SBP = systolic blood pressure; SD = standard error;
SE = standard error.
Note: fRed text: Studies published since the 2013 Pb ISA, Black text: Studies included in the 2013 Pb ISA.
Effect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb. If the Pb biomarker is log-
transformed, effect estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker
level. Effect estimates are assumed to be linear within the evaluated interval. Categorical effect estimates are not standardized.
Figure 4-1 Association between biomarkers of Pb exposure and blood
pressure.
Many nationally representative cross-sectional studies evaluated the association between
increases in BLLs and changes in either SBP or DBP using continuous Pb biomarkers. Generally,
increases in BLLs were concurrently associated with higher SBP and DBP (Huang. 2022; Qu et al.. 2022;
Teve et al.. 2020; Lee et al.. 2016a; Hara et al.. 2015; Hicken et al.. 2013; Scinicariello et al.. 2011).
However, some nationally representative studies noted null associations for SBP, but positive associations
for DBP (Obcng-Gvasi et al.. 2018; Bushnik et al.. 2014; Zota et al.. 2013a). while others noted positive
associations for SBP and null associations for DBP (Everson et al.. 2021). Studies containing the
necessary information to standardize effect estimates to a 1 (ig/dL increase in blood Pb or a 10 |ig/g
increase in bone represent similar trends (Figure 4-1) and conclusions as studies that did not contain the
necessary information needed for standardization (Figure 4-1 and Table 4-3). Specifically, in a
KNHANES (2008-2013) analysis, Lee et al. (2016a) reported 0.71 mmHg higher DBP with each
doubling of blood Pb (95% CI: 0.29, 1.13 mmHg) and a similar association with SBP (0.73 mmHg [95%
CI: 0.09, 1.36 mmHg]). In an NHANES (1999-2006) analysis Scinicariello et al. (2011) indicated higher
SBP (1.07 mmHg [95% CI: 0.384, 1.76 mmHg]) and higher DBP (0.71 mmHg [95% CI: 0.18,
1.24 mmHg]) for a twofold higher BLL. In contrast, in an analysis of more recent NHANES cycles
(2007-2010), Obeng-Gvasi et al. (2018) noted a 0.268 mmHg higher DBP (95% CI: 0.079,
0.458 mmHg), but reported a null association between ln-Pb and SBP (0.052 mmHg [95% CI: -0.233,
0.458 mmHg]) (Table 4-3).
Several smaller cross-sectional studies have also examined the relationship between Pb
biomarkers and BP (Yan et al.. 2022; Xu et al.. 2021; Chung et al.. 2020; Wang et al.. 2020; Guo et al..
2019; Lopes et al.. 2017b; Gambelunghe et al.. 2016; Ettinger et al.. 2014). These studies tended to
support the larger nationally representative studies. Several studies indicated positive associations both
SBP and DBP (Yan et al.. 2022; Chung et al.. 2020; Wang et al.. 2020; Gambelunghe et al.. 2016). For
example, a moderately sized study (n = 770) in Taiwan noted higher SBP (1.34 mmHg [95% CI: 0.34,
2.52 mmHg]) and DBP (0.69 mmHg [95% CI: 0.01, 1.37 mmHg]) per 1 (ig/dL higher blood Pb (Chung et
al.. 2020). Additionally, a large cohort (with a cross-sectional component) in Malmo, Sweden (n = 4,452)
assessed BP and BLLs in the early nineties (1991-1994). This population was likely exposed to
historically high levels of Pb in the environment. The fully adjusted model indicated higher SBP
(1.8 mmHg [95% CI: 0.52, 3.08 mmHg]) and DBP (1.4 mmHg [95% CI 0.57, 2.54 mmHg]) when
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comparing the highest quartile of BLLs (mean 4.7 (ig/dL) with the lower three quartiles (mean 1.5-
2.8 (ig/dL) (Gambelunghe et al.. 2016). Yan et al. (2022) (n = 2.504) cross-sectionally evaluated a Haitian
population with relatively higher BLLs (geometric mean [GM]: 4.73 (ig/dL). This study had a high limit
of detection (3.3 (.ig/dL). however, and -30% of the study population had BLLs below the limit of
detection. Yet, this study indicated positive associations between blood Pb and SBP (2.42 mmHg [95%
CI: 0.36, 4.49]) and DBP (1.96 mmHg [95% CI: 0.56, 3.37]) when comparing the highest quartile (6.5-
58.2 (ig/dL) with the lowest (<3.3 (ig/dL).
In addition, a moderately sized study (n = 816) among older adults (aged 40-75) living in rural
southwest China compared the associations between BLL quartiles and BP measurements among those
subsisting off rice and vegetables grown in a polluted region (Cd) concentration >0.2 mg/kg) with the
associations among those in an unpolluted region (Cd <0.05 mg/kg) (Wang et al.. 2020). In the polluted
region, this study reported positive associations with both SBP and DBP when the highest BLL quartile
(>4.6 (ig/dL) was compared with the lowest BLL quartile (<2.1 (ig/dL) (Figure 4-2). In contrast, there was
no relationship observed in the unpolluted region (Figure 4-3). The authors of this study hypothesize that
this discrepancy in association between polluted and unpolluted regions may be due in part to differences
in mean BLLs in the polluted (3.5 (ig/dL) and unpolluted (2.6 (ig/dL) areas, or that more pollution may
modify the association between blood Pb and BP, in addition to the small sample size in the unpolluted
area (n = 214) compared with the polluted area (n = 602).
Q1(<20 SO)
Q2(20 80-34 10)
Pb 03(34 10-46 40)
Q4(>46 40)
Hypertension
p-trend
SBP
p-trend
DBP
/Mrend
0.438
i
O.OOS
<0.001
.
A
I A !
2 3 4
-5
5 10
-5 0 5 10
Odds Ratios
Coefficients
Coefficients
DBP = diastolic blood pressure; Pb = lead; Q = quartile; SBP = systolic blood pressure.
Source: Adapted from Wang et al. (2020).
Figure 4-2 Association between blood Pb level quartiles and systolic blood
pressure, diastolic blood pressure, and hypertension, polluted
region of rural southwest China.
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Q1(<16 00)
02(16 00-2375)
Pb 03(23 75-37 98)
Q4(»37 98)
Hypertension
1
p trend SBP
p-trend DBP
0.567
3 6 9
Odds Ratios
0,76V
p trend
0.524
Coefficients
-S 0 5
Coeffiaents
DBP = diastolic blood pressure; Pb = lead; Q = quartile; SBP = systolic blood pressure.
Source: Adapted from Wang et al. (2020).
Figure 4-3 Association between blood Pb level quartiles and systolic blood
pressure, diastolic blood pressure, and hypertension, unpolluted
region of rural southwest China.
In contrast, a study in Cambe, Brazil (n = 948) indicated a null association between BLLs and
SBP. However, when comparing the 90th percentile (6.03 (ig/dL) with the 10th percentile (0.74 (.ig/dL).
DBP was 0.005 mmHg higher (95% CI: 0.002, 0.008 mmHg) per 1 (ig/dL increase in blood Pb
concentration (Lopes et al.. 2017b). Additionally, a recent study cross-sectionally evaluated the
association between blood Pb and BP among participants in the Gulf Long-Term Follow-up (GuLF) study
(Xu et al.. 2021). The GuLF study is a longitudinal cohort of individuals involved in the 2010 Deepwciter
Horizon oil spill. Baseline blood Pb and BP measurements were obtained between 2011 and 2013. BLLs
within this study were low overall (quartile 1: 0.06 (ig/dL, quartile 4: 0.27 (.ig/dL). This study indicated
null associations between the highest quartile and the lowest quartile for SBP (-0.96 [95% CI: -4.13,
2.22]) and DBP (-0.01 mmHg [95% CI: -2.21, 2.10]).
Additionally, a smaller number of cross-sectional studies evaluated either SBP or DBP
categorically. Typically, these studies dichotomized either SBP or DBP at a particular clinically relevant
threshold prior to conducting categorical statistical analyses. The results of these studies were more mixed
compared with the results presented using continuous BLLs, presented above. For example, in a small
study (n = 150) Ettinger et al. (2014) evaluated the association between BLLs and high SBP
(>130 mmHg) and high DBP (>85 mmHg) among young adults (aged 25-45) of African descent. This
study yielded null results for both high SBP (>130 mmHg) (OR: 1.69 [95% CI: 0.55, 5.15]) and high
DBP (>85 mmHg) (OR: 2.20 [95% CI: 0.59, 8.16]) when comparing blood Pb values above and below
the median (1.66 (ig/dL). In contrast, a different larger study among young adults (aged 18-44)
(n = 7,730) indicated higher odds of SBP >120 mmHg (OR: 1.21 [95% CI: 1.07, 1.38]) and DBP
>80 mmHg (OR: 1.32 [95% CI: 1.10, 1.58]) when comparing BLLs above and below 5 (ig/dL (Obeng-
Gvasi. 2019). These cross-sectional results were similar to the results generated from studies evaluating
concurrent BLLs and hypertension (see Section 4.3.1.2).
While most cross-sectional studies evaluated the association between concurrent BLLs and SBP
and DBP, some studies assessed in the 2013 Pb ISA also considered concurrent bone Pb measurements
and BP. Bone Pb tends to represent cumulative or long-term exposure to Pb, whereas BLLs are
representative of recent exposure. Several analyses previously presented in the 2013 Pb ISA indicated
4-14
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mixed results for the association between bone Pb levels and SBP and DBP, although associations were
generally positive. For example, (Elmarsafawv et al.. 2006) evaluated whether calcium intake affects the
relationship between SBP and bone Pb levels in a cross-sectional analysis of the NAS cohort.
(Elmarsafawv et al.. 2006) reported higher SBP for each 10 (ig/g increase in bone Pb level in both high
(>800 mg/day) and low calcium (<800 mg/day) groups. However, the association with bone Pb was
substantially larger in the low calcium group (4.00 mmHg [95% CI: 1.05, 6.95]) than in the high calcium
group (1.90 mmHg [95% CI: 0.10, 3.70]). Although these results are not presented in Figure 4-1, they are
comparatively larger than results presented for SBP and BLLs. No recent studies evaluated the
association between bone Pb levels and BP.
In addition to the numerous cross-sectional studies previously mentioned, several recent studies
longitudinally evaluated the relationship between biomarkers of Pb exposure and BP measurements.
Bulka et al. (2019) evaluated a small Bangladeshi cohort (n = 255) with baseline BLLs (median:
8.5 (ig/dL) measurements between April 2006 and August 2009, from an arsenic-endemic area. Residents
in this area are chronically exposed to high levels of Pb in the air, water, and other industrial sources. BP
was assessed biennially for a total of 6 years. This study indicated that SBP was increased in the highest
quartile of baseline blood Pb compared with the lowest quartile, corresponding to a 1.16 mmHg annual
increase (95% CI: 0.21, 2.11 mmHg). Results for DBP (0.53 mmHg [95% CI: -0.10, 1.16 mmHg]) and
PP (0.63 mmHg [95% CI: -0.08, 1.34]) were smaller in magnitude, compared with SBP when comparing
the highest quartile of BLLs to the lowest quartiles. All analyses considered several appropriate
confounders, in addition to urinary arsenic (creatinine standardized). While there was an annual increase
in the association between blood Pb and SBP, BP measurements remained stable across visits, but
antihypertensive medication use increased from 7.5% at baseline to 15.3% at the last visit, which was
controlled for as a confounder in all statistical models. A longitudinal study in Belgium(Yu et al.. 2020)
(n = 267) evaluated the association between baseline BLLs (collected between 1985 and 2005) and BP
measured an average of 9.4 years following blood Pb measurement (Yu et al.. 2020). For each doubling
of BLLs there were null associations between peripheral SBP (2.41 mmHg [95% CI: -0.38,
5.20 mmHg]), DBP (0.50 mmHg [95% CI: -1.07, 2.07 mmHg]), and PP (1.91 mmHg [95% CI: -0.32,
4.14 mmHg]). Similarly, the association between a doubling of BLLs and central SBP, DBP, and PP were
also null. Overall, associations from these studies remained stable even after controlling for Cd at baseline
and considering the high endemic levels of arsenic in the Bangladeshi cohort.
Several studies also assessed PP in addition to SBP and DBP. To reiterate, PP is the force the
heart requires to contract and is calculated by subtracting DBP from SBP. Overall, there was a null
relationship between BLLs and PP in both cross-sectional (Hara et al.. 2015; Scinicariello et al.. 2010
Perlstein. 2007. 194019) and cohort analyses (Yu et al.. 2020; Bulka et al.. 2019). However, the
relationship between bone Pb levels and PP was positive in cross-sectional analyses (Zhang et al.. 2010;
Perlstein et al.. 2007). Hara et al. (2015) additionally evaluated MAP, which is the average BP during a
single cardiac cycle. This study indicated an increase in MAP associated with BLLs.
4-15
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The 2013 Pb ISA included two different meta-analyses focused on the relationship between Pb
exposure biomarkers and BP changes or hypertension status. Nawrot et al. (2002) included over 30 cross-
sectional and prospective studies on BLLs and BP, including >58,000 adults. This meta-analysis
concluded that each doubling of concurrent BLLs was associated with a 1 mmHg increase in systolic BP
and a 0.6 mmHg increase in diastolic BP. Furthermore, Navas-Acien et al. (2008) conducted a similar
meta-analysis based on bone Pb measurements (three prospective, five cross-sectional). The pooled
estimate from the cross-sectional studies indicated an increase in SBP of 0.26 mmHg (95% CI: 0.02,
0.50) per 10 (ig/g tibia Pb. When considering hypertension, pooled results indicated increased odds of
hypertension (OR: 1.04 [95% CI: 1.01, 1.07]) per 10 (ig/g increase in tibia Pb and 1.04 (95% CI: 0.96,
1.12) per 10 |ig/g increase in patella Pb.
4.3.1.1.1 Effect Measure Modification
Several recent studies went beyond only evaluating the direct association between BLL and BP,
but also evaluated effect measure modification (EMM) by several different variables, including race
(Huang. 2022; Teve et al.. 2020; Hara et al.. 2015; Hicken et al.. 2013; Hicken et al.. 2012; Scinicariello
et al.. 2011). sex (Gambclunghc et al.. 2016; Hara et al.. 2015; Bushnik et al.. 2014; Hicken et al.. 2013;
Hicken et al.. 2012; Scinicariello et al.. 2011). age (Huang. 2022; Obeng-Gvasi. 2019; Gambelunghe et
al.. 2016; Bushnik et al.. 2014). stress/depression (Hicken et al.. 2013; Zota et al.. 2013a). genetic
variations (Jhun et al.. 2015). and smoking (Gambelunghe et al.. 2016). These analyses can help to further
highlight specific subgroups of the population that may have an increased risk of elevated BP associated
with Pb biomarkers of exposure.
Race was a common measure to evaluate differential effects of biomarkers of Pb exposure and
BP. An NHANES (1999-2016) study indicated that both SBP and DBP were statistically significantly
higher among non-Hispanic white individuals (SBP: 0.34 mmHg [95% CI: 0.11, 0.57 mmHg], DBP:
0.38 mmHg [95% CI: 0.19, 0.57 mmHg]) and non-Hispanic Black individuals (SBP: 0.67 mmHg [95%
CI: 0.29, 1.05 mmHg], DBP: 0.36 mmHg [95% CI: 0.06, 0.66 mmHg]), with each 1 (ig/dL increase in
BLLs compared to other races evaluated in the studv(Teve et al.. 2020). As described in the 2013 Pb ISA,
Scinicariello et al. (2010) used NHANES III (1988-1994) to evaluate BP and BLLs by race/ethnicity.
This study indicated a higher SBP (1.615 mmHg [95% CI: 1.007, 2.223 mmHg]) and DBP (1.261 mmHg
[95% CI: 0.716, 1.805 mmHg]) among non-Hispanic Black individuals per 1 (ig/dL higher blood Pb,
compared with non-Hispanic white and Mexican-American individuals (Figure 4-1, Table 4-3).
Evaluation of effect modification by sex was less common, and the results were less consistent
than for race. The Malmo Diet and Cancer Study evaluated the relationship between BLLs and changes in
BP stratified by sex (Gambelunghe et al.. 2016). In this study, sex did not modify the positive association
between BLLs and SBP or DBP increases. Additionally, an NHANES (2003-2010) analysis reported
4-16
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higher SBP and DBP for each doubling of BLLs among both sexes (Haraet al.. 2015) (Figure 4-4,
Table 4-3).
Several studies evaluated the intersectionality of sex and race as potential modifiers of the
association between BLLs and changes in BP. In a cross-sectional study using NHANES (2003-2010),
Hara et al. (2015) evaluated SBP and DBP stratified by both sex and race (Figure 4-4). This study
indicated that qualitatively, compared with white females, Black females experienced higher SBP with
each doubling of BLLs. White females had higher DBP, compared with Black females, however.
Compared with white males, Black and Hispanic males had higher SBP with each doubling of BLLs.
White and Black men had similar associations between higher DBP and higher BLLs (Figure 4-4,
Table 4-3). Scinicariello et al. (2011) used NHANES (1999-2006) to evaluate sex and racial disparities
for changes in BP and BLLs (Figure 4-5). This study reported the highest SBP among Black males
(2.40 mmHg [95% CI: 0.91, 3.69 mmHg]) and Black females (2.40 mmHg [95% CI: 0.17, 4.63 mmHg])
associated with a doubling of BLLs, compared with other races. Conversely, Mexican-American males
had lower DBP associated with a doubling of BLLs (-1.34 mmHg [95% CI: -2.63, -0.05 mmHg]).
Huang (2022) also conducted an analysis using NHANES (1999-2006). This study generally noted
similar positive associations with higher SBP and DBP among non-Hispanic white males and females and
non-Hispanic Black males and females; however, there were null associations for Mexican-American and
other Hispanic males and females for both SBP and DBP.
4-17
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Population
SBP:
All
Women
Black
Hispanic
White
Men
Black
Hispanic
White
DBP:
All
Women
Black
Hispanic
White
Men
Black
Hispanic
White
PP:
All
Women
Black
Hispanic
White
Men
Black
Hispanic
White
MAP:
All
Women
Black
Hispanic
White
Men
Black
Hispanic
White
Pb distribution
1.37 (0.88-2.10)
1.21 (0.80-1.78)
1.22 (0.80-1.86)
1.86(1.20-2.85
1.94 (1.25-2.83)
1.73(1.16-2.57)
1.37(0.88-2.10) *
1.21 (0.80-1.78) *-+-
1.22 (0.80-1.86)
1.86(1.20-2.85
1.94 (1.25-2.83)
1.73 (1.16-2.57)
1.37 (0.88-2.10)
1.21 (0.80-1.78)
1.22 (0.80-1.86)
1.86 (1.20-2.85
1.94 (1.25-2.83)
1.73(1.16-2.57)
1.37(0.88-2.10)
1.21 (0.80-1.78)
1.22 (0.80-1.86)
1.86(1.20-2.85
1.94 (1.25-2.83)
1.73 (1.16-2.57)
EE
LCL
UCL
0.76
0.38
1.13
0.58
0.01
1.17
* 1.18
-0.19
2.55
¦ 0.56
-0.57
1.69
0.61
-0.18
1.40
0.79
0.30
1.27
* 1.61
0.45
2.76
* 0.95
0.05
1.84
0.65
-0.03
1.32
0.43
0.18
0.68
0.43
0.07
0.80
0.52
-0.34
1.37
-0.13
-0.81
0.55
0.73
0.23
1.24
0.47
0.13
0.81
0.81
-0.04
1.66
-0.03
-0.64
0.58
0.70
0.24
1.17
0.33
-0.02
0.67
0.15
-0.37
0.67
0.64
-0.56
1.84
- 0.69
-0.29
1.67
-0.11
-0.84
0.61
0.32
-0.13
0.77
-~0.81
-0.27
1.88
0.98
0.14
1.82
-0.06
-0.68
0.57
—I 1—
0.00 0.50
mmHg (per doubling of blood Pb)
0.54
0.29
0.79
0.48
0.10
0.86
0.73
-0.16
1.62
0.10
-0.62
0.82
0.69
0.18
1.21
0.57
0.24
0.91
¦* 1.08
0.26
1.90
0.30
-0.30
0.89
0.68
0.23
1.14
1.00
1.50
DBP = diastolic blood pressure; EE = effect estimate; MAP = mean arterial pressure; Pb = lead; PP = pulse pressure; SBP = systolic
blood pressure; LCL = lower confidence limit; UCL = upper confidence limit.
Note: Pb distribution presented as geometric mean (IQR).
Source: Hara et al. (2015).
Figure 4-4 Effect measure modification by sex and race for blood pressure
(systolic and diastolic) and a doubling of blood Pb levels,
National Health and Nutrition Examination Survey (2003-2010).
4-18
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Population
SBP:
All
White:
Men
Women
Black:
Men
Women
Pb distribution
2.2 (0.03)
1.55 (0.02)
2.44 (0.05)
1.81 (0.06)
Mexican-American:
Men
Women
DBP:
All
White:
Men
Women
Black:
Men
Women
2.47 (0.06)
1.56 (0.04)
2.2 (0.03)
1.55 (0.02)
2.44 (0.05)
1.81 (0.06)
Mexican-American:
Men
Women
PP:
All
White:
Men
Women
Black:
Men
Women
2.47 (0.06)
1.56 (0.04)
2.2 (0.03)
1.55 (0.02)
2.44 (0.05)
1.31 (0.06)
Mexican-American:
Men 2.47 (0.06)
Women 1.56 (0.04)
T
T
"T
—I
-1.00 0.00 0.50 1.00
mmHg (per doubling of blood Pb)
T
T
1.50 2.00
EE LCL
1.07 0.38
0.87 -0.17
0.89 -0.19
~2.30 0.91
~2.40 0.17
0.10 -1.27
-0.03 -1.28
0.90 0.02
0.95 0.21
—I 1
2.50 3.00
UCL
1.76
1.91
1.97
3.69
4.63
1.47
1.22
0.71 0.18 1.24
—~2.75 1.14
0.30 -1.29
-1.34 -2.63
-0.74 -1.60
1.73
1.69
4.36
1.89
-0.05
0.12
0.37 -0.30 1.04
-0.02 -1.10
-0.03 -1.25
-0.42 -2.24
-2.21 -0.08
1.42 0.05
0.70 -0.53
1.06
1.19
1.40
4.50
2.79
1.93
DBP = diastolic blood pressure; EE = effect estimate; LCL = lower confidence limit; PP = pulse pressure; SBP = systolic blood
pressure; UCL = upper confidence limit.
Note: Pb distribution presented as mean (SE).
Source: Scinicariello et al. (2011).
Figure 4-5 Effect measure modification by sex and race for blood pressure
(systolic, diastolic, and pulse pressure) and a doubling of blood
Pb level, National Health and Nutrition Examination Survey (1999-
2006).
Additionally, using NHANES (2001-2008) Hicken et al. (2012) indicated differences in SBP,
DBP, and PP when comparing white and Black males and females (Figure 4-6). The associations between
BLLs and SBP, DBP, and PP were consistently higher among Black females compared with white
females. Furthermore, this discrepancy by race and sex was also altered by educational attainment
(Figure 4-7) and family poverty (Figure 4-8).
4-19
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Population
SBP:
White Men
< High School
£ High School
Poor
Non-Poor
White Women
< High School
a High School
Poor
Non-Poor
Black Men
< High School
a High School
Poor
Non-Poor
Black Women
< High School
2 High School
Poor
Non-Poor
DBP:
White:
Men
Women
Black:
Men
Women
PP:
White:
Men
Women
Black:
Men
Women
Blood Pb
1.7(1.7)
1.2 (1.2)
1.9 (1.8)
1.4(1.3)
EE LCL UCL
1.7(1.7)
1.2 (1.2)
1.9 (18)
1.4(1.3)
1.7 (1.7)
1.2 (1.2)
1.9 (1.8)
1.4(1.3)
0.3
0.3
0.3
0.6
0.3
0.8
1.0
-0.88
-2.25
-2.25
-1.16
-1.07
-0.96
-2.14
0.3 -1.86
0.7 -1.06
1.1
2.8
4.1
2.2
4.8
0.4
4.0
1 6.7
2.9
" 4.3
•3.7
-0.47
1.43
1.55
0.24
4.21
-1.36
1.84
2.58
-0.24
1.36
-0.02
1.3
2.1
1.48
2.85
2.85
2.36
1.67
2.56
4.14
2.46
2.46
2.67
4.17
6.65
4.16
5.39
2.16
6.16
10.82
6.04
7.24
7.42
0.6 -0.18 1.38
1.3 0.32 2.28
1.5 -0.07 3.07
1.9 0.53 3.27
-0.3 -1.28 0.68
-0.5 -1.87 0.87
-0.27
0.34
2.87
3.86
1 1 1 1—
0.00 1.00 2.00 3.00
mmHg (per doubling of blood Pb)
4.00
5.00
6.00
7.00
DBP = diastolic blood pressure; EE = effect estimate; LCL = lower confidence limit; Pb = lead; PP = pulse pressure; SBP = systolic
blood pressure; UCL = upper confidence limit.
Note: Pb distribution presented as mean (median).
Source: Hicken et al. (2012).
Figure 4-6 Effect measure modification by sex and race for blood pressure
(systolic, diastolic, and pulse pressure) and a doubling of blood
Pb levels, National Health and Nutrition Examination Survey
(2001-2008).
4-20
-------
a
White, high school
— Black, < high school Black, £ high school
b
Q-
11S -| 1 1 1 1 1 1 1 1 1
-0.53 -0.33 -0.12 0.00 0.18 0.34 0.48 0.70 0.97
Blood Lead, jig/dL
SBP = systolic blood pressure.
Note: Association between SBP and log-transformed BLL by educational attainment for men (a) and women (b).
Source: Hicken et al. (2012).
Figure 4-7 Effect measure modification between blood Pb levels, race, and
education level, National Health and Nutrition Examination Survey
(2001-2008).
4-21
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White, PIR < 1.85 (poor) White, PIR >1.85 (nonpoor)
Black, PIR < 1.85 (poor) ¦¦¦¦« Black, PIR >1.85 (nonpoor)
130
Di 128
X
~U
qj
T3
ai
" 126
124
122
120
118
1 1 1 1 1 1 1 1 1
-0.53 -0.33 -0.12 0.00 0.18 0.34 0.48 0.70 0.97
Blood Lead, ng/dL
SBP = systolic blood pressure, PIR = poverty-income ratio.
Note; Association between SBP and log-transformed BLL by poverty level for men (a) and women (b).
Source: Hicken et al. (2012).
Figure 4-8 Effect measure modification between blood Pb levels, race, and
poverty level, National Health and Nutrition Examination Survey
(2001-2008).
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Another NHANES (2005-2008) analysis further evaluated EMM by racial differences and
depressive symptoms on the effects of a doubling of BLLs on BP (Hicken et al.. 2013). First, the
association between higher SBP and a doubling of BLLs was larger among Black participants (3.2 mmHg
[95% CI: 1.5, 5.0 mmHg]) than white participants (1.0 mmHg [95% CI: -0.3, 2.4 mmHg]). However,
higher DBP was similar when comparing Black and white participants. This study further evaluated
potential EMM by considering depressive symptoms, defined using the Patient Health Questionnaire
(PHQ-9), which may indicate psychosocial stress. The PHQ-9 score was parsed into low (score <3) and
high (score >3). The association between BLLs and BP (both SBP and DBP) was greater among those
with high PHQ-9 scores. High psychosocial stress (PHQ-9 score >3) particularly modified the association
between blood Pb and BP among Black individuals, compared with white individuals (Table 4-3).
Specifically, a doubling of BLLs was associated with 5.6 mmHg (95% CI: 2.0, 9.2 mmHg) higher SBP
among Black individuals with high levels of depression (PHQ-9 score >3), compared with only
1.2 mmHg (95% CI: -0.5, 2.9 mmHg) higher SBP among white individuals with high levels of
depression (PHQ-9 score >3).(Zota et al.. 2013a)
Several studies also evaluated if age was an effect modifier of the relationship between
biomarkers of Pb exposure and changes in BP. Specifically, an NHANES (2009-2016) analysis evaluated
the odds of the associations between BLLs and higher SBP (>120 mmHg) or DBP (>80 mmHg) for
middle-aged (46-65 years) and young (aged 18-44 years) adults (Obeng-Gvasi. 2019). This study
demonstrated similar odds of higher SBP for middle-aged adults (OR: 1.32 [95% CI: 1.14, 1.52]) as with
young adults (OR: 1.21 [95% CI: 1.07, 1.38]) when comparing BLLs above and below 5 (ig/dL. The
association between BLLs and higher DBP was also similar in middle-aged (OR: 1.16 [95% CI: 0.98,
1.38]) and young (OR: 1.32 [95% CI: 1.10, 1.58]) adults. The young adults included in this analysis were
likely not exposed to air emissions associated with leaded gasoline in the past, and therefore can help
disentangle the effects of past high Pb exposures on CVD health endpoints. Additionally, the Malmo Diet
and Cancer Study also considered both sex and age as potential effect modifiers when evaluating
associations between BLLs and changes in BP (Gambelunghe et al.. 2016). This study noted marginally
increased associations between BLLs and higher SBP among adults aged >57 years (2.4 mmHg [95% CI:
1.20, 3.60 mmHg]) compared with adults <57 years (1.3 mmHg [95% CI: -0.55, 3.15]), when comparing
the highest blood Pb quartile (4.7 (ig/dL) with the lowest three quartiles (range 1.5-2.8 (.ig/dL). There
were no differences, however, in the association between BLLs and higher DBP by age. However, this
cohort was likely exposed to air emissions associated with leaded gasoline in the past. In addition, a
cross-sectional study, using the Canadian Health Measures Survey (2007-2011), (Bushnik et al.. 2014)
demonstrated a steep increase in SBP and DBP associated with BLLs, up to 3 (ig/dL, especially among
middle-aged adults (40-54 years) and men. Specifically, this study indicated for each 1 (ig/dL of BLL
would correspond with a 1-2 mmHg higher SBP and a 2-3 mmHg higher DBP (Figure 4-9, Figure 4-10).
4-23
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Systolic blood pressure (mm Hg)
3 4
Blood lead level (pg/dL)
¦ Model 1 (40 to 79 years)
¦ Model 3 (55 to 79 years)
Model 5 (Women, 40 to 79 years)
- Model 2 (40 to 54 years)*
¦¦ Model 4 (Men, 40 to 79 years)
'significant association between blood Pb level and systolic blood pressure (p < 0.05).
BMI = body mass index; HDL = high-density lipoprotein.
Source: Bushnik et al. (2014).
Figure 4-9 Effect measure modification by sex and age of the relationship
between blood Pb levels and systolic blood pressure, Canadian
Health Measures Survey.
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Diastolic blood pressure (mm Hg)
80-
79-
78-
77-
76-
75-
74-
73-
72-
71-
70-
*significant association between blood Pb level and diastolic blood pressure (p < 0.05)
BMI = body mass index; HDL = high-density lipoprotein.
Source: Bushnik et al. (2014).
Figure 4-10 Effect measure modification by sex and age of the relationship
between blood Pb levels and diastolic blood pressure, Canadian
Health Measures Survey.
Certain genetic polymorphisms can be important in assessing the risk of increased BP as a result
of elevated levels of biomarkers of Pb exposure, and therefore can be an important effect modifier to
evaluate. In a longitudinal analysis of the NAS cohort, Jhun et al. (2015) evaluated potential EMM by
vitamin D receptors (VDR) between PP and bone level and BLL. Genetic variations in VDR genes can
potentially influence the accumulation, absorption, and retention of Pb in the body. After the initial
baseline bone Pb, blood Pb, and BP measurements, PP was reassessed every 3-5 years. At the initial visit,
an IQR increase in either tibia or patella Pb level was associated with an increased PP among those with
the variant (opposed to ancestral) genotype (single nucleotide polymorphisms [SNPs] in Bsml, Taql,
Apal, or Fokl). Although there was an association with PP and tibia Pb levels by VDR genotype at
baseline, this relationship appeared to diminish with time (Figure 4-11). However, the three-way
3 4
Blood lead level (|jg/dL)
¦ Model 1 (40 to 79 years)*
¦Model 3 (55 to 79 years)
Model 5 {Women, 40 to 79 years)
Model 2 (40 to 54 years)*
-—-Model 4 (Men, 40 to 79 years)*
4-25
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interaction terms between bone Pb levels, VDR receptor type, and time since baseline, used to further
assess EMM, was almost zero, indicating that VDR consistently modifies the association between bone
Pb levels and PP. In addition to genetic polymorphisms in VDR, the 2013 Pb ISA also evaluated studies
that assessed other genetic factors that may increase susceptibility to Pb. Specifically, Scinicariello et al.
(2010) used NHANES III (1988-1994), to stratify by S-aminolevulinic acid dehydratase (ALAD) status.
A critical mechanism of Pb toxicity is its ability to interact and inhibit key enzymes, such as ALAD, in
the heme biosynthetic pathway. This study indicated that non-Hispanic white carriers of the ALAD 2
polymorphism had higher measures of SBP and DBP associated with BLLs. However, in a South Korean
occupational study, BLLs were associated with higher SBP only, and there was no evidence of EMM by
either VDR or ALAD (Weaver et al.. 2008). In another evaluation of the MAS, Zhang et al. (2010)
examined changes in the hemochromatosis gene (HFE), which can promote excessive iron absorption and
is thought to also alter Pb biomarker concentrations. Two mutations to HFE (C282Y and H63D) were
examined within this older population. This study suggested that those with the H63D mutation were
more likely to have an increase in PP with a 10 (.ig/g increase in tibia (2.54 mmHg [95% CI: 0.12,
4.96 mmHg]) and patella (2.23 mmHg [95% CI: 0.23, 4.23 mmHg]) Pb levels. Taken together, certain
genetic polymorphisms appear capable of predisposing some groups to greater effects on BP related to
biomarkers of Pb exposure.
At baseline
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Tibia lead levels (ng/g)
-Ancestral - Variant
BMI = body mass index; VDR = vitamin D receptor.
Source! Jhun et al. (2015).
Figure 4-11 Effect measure modification by vitamin D receptor variant for the
association between pulse pressure and tibial Pb levels,
Normative Aging Study cohort.
The Malmo Diet and Cancer Study further evaluated EMM by smoking status for the association
between BLLs and BP (Gambelunghe et al.. 2016). The cross-sectional component of the study indicated
that smokers (ever-smokers) had 3.9 mmHg (95% CI: 1.59, 6.21 mmHg) higher SBP, compared with only
0.6 mmHg (-1.46, 2.66 mmHg) among never-smokers when comparing the highest quartile of BLLs
4-26
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(mean 4.7 (ig/dL) with the lower three quartiles (mean 1.5-2.8 (ig/dL). Similarly, smokers had a
1.6 mmHg (95% CI: 0.65, 2.54 mmHg) higher DBP, compared with 1.1 mmHg (95% CI: -0.05,
2.25 mmHg) higher DBP among never-smokers.
4.3.1.2 Hypertension
Fewer recent studies evaluated the relationship between biomarkers of Pb exposure and
hypertension. Study-specific details, including blood and bone Pb levels, study population characteristics,
potential confounders, and select results from these studies are highlighted in Table 4-4 and Figure 4-12.
Studies in Figure 4-4 are standardized to represent the risk of prevalent or incident hypertension
associated with a 1 (ig/dL increase in BLL or a 10 |ig/g increase in bone Pb level. Study details shown in
Table 4-4 include standardized results as well as results that could not be standardized based on the
information provided in each paper. Generally, hypertension refers to chronic BP readings of
>140 mmHg for SBP and >90 mmHg for DBP, while prehypertension, typically thought of as a precursor
to chronic hypertension, is usually defined as SBP 120-139 mmHg or DBP 80-89 mmHg. However,
some studies may choose to define hypertension, or prehypertension, differently. Some cross-sectional
studies evaluated associations between biomarkers of Pb exposure and prevalent or preexisting
hypertension (Huang. 2022; Qu et al.. 2022; Xu et al.. 2021; Teve et al.. 2020; Wang et al.. 2020; Choi et
al.. 2018; Lopes et al.. 2017a; Hara et al.. 2015; Bushnik et al.. 2014). However, other studies specifically
evaluated prehypertension, or SBP or DBP values that approach a predefined clinical definition of
hypertension (Ou et al.. 2022; Lee et al.. 2016b; Lee et al.. 2016a). Additionally, longitudinal studies
examined associations between baseline Pb biomarkers and incident, or newly developed hypertension
(Gambelunghe et al.. 2016) whereas other studies evaluated associations with hypertension that may not
respond to medication (resistant) or completely untreated (uncontrolled) hypertension using NHANES
(Miao et al.. 2020) or the NAS cohort (Zheutlin et al.. 2018).
4-27
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Reference Study Design
OR:
tAlmeida Lopes et al, 2017 Cross-sectional
Population
Adults >40 Cambe, Brazil
Pb distribution Pb biomarker
Geometric mean: 1.97 (95%CI:1.90-2.04) Blood
tChoi et al, 2018
Cross-sectional
KNHANES
Curry Intake
Non-Curry Intake
Mean (SE): 2.01 (0.025)
Blood
TMiao et al, 2020
Cross-sectional NHANES
Mean (SE):
Male: 1.50(0.02)
Female: 1.07 (1.01)
Uncontrolled Hyptertension
vs Non-hypertension
Male
Female
Uncontrolled Hypertension
vs Controlled Hypertension
Male
Female
Uncontrolled Hypertension vs
Controlled and Non-Controlled Hypertension
Male
Female
TTeye et al, 2020
THuang et al, 2022
Cross-sectional
Cross sectional
NHANES
NHANES
All
Women
Men
Mexican American
Other Hispanic
Non-Hispanic White
Non-Hispanic Black
Other Race
Mean (SD): 1.73 (1.71)
Blood
Blood
Elmarsafawy et al, 2006 Cross-sectional
RR:
tZheutlin et al, 2018
NAS Men
Low Calcium
High Calcium
NAS Men, Resistant hypertension
Mean(SD)
21.6(12.0)
31.7 (18.3)
6.6 (4.3)
21.6(12.0)
31.7 (18.3)
6.6 (4.3)
Median IQR
20.0 (13.0-28.5)
27.0 (18.0—40.0)
5 (3.4-5.0)
Tibia
Patella
Blood
Tibia
Patella
Blood
Tibia
Patella
Blood
0.75 1.00 1.25
Change in BP (mmHg 95% CI) per 1 ug/dL increase in blood Pb or 10 ug/dL increase in bone Pb
BP = blood pressure; CI = confidence interval; IQR = interquartile range; KNHANES = Korea National Health and Nutrition
Examination Survey; NAS = Normative Aging Study; NH = non-Hispanic; NHANES = National Health and Nutrition Examination
Survey; OR = odds ratio; Pb = lead; RR = relative risk; SD = standard error; SE = standard error.
Note: fRed text: Studies published since the 2013 Pb ISA, Black text: Studies included in the 2013 Pb ISA.
Effect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb. If the Pb biomarker is log-
transformed, effect estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker
level. Effect estimates are assumed to be linear within the evaluated interval. Categorical effect estimates are not standardized.
Figure 4-12 Associations between biomarkers of Pb exposure and
hypertension.
Cross-sectional studies identified positive associations between BLLs and prevalent hypertension
but were not statistically significant. Wang et al. (2020) (n = 816) indicated no association between BLLs
and hypertension prevalence among an older adult Chinese population (Figure 4-2 and Figure 4-3).
Similarly, Bushnik et al. (2014). (n = 4,550) also indicated no association between BLLs and
hypertension prevalence among participants of the Canadian Health Measures Survey. Studies evaluating
4-28
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NHANES (1999-2016) (Teve et al.. 2020). NHANES (1999-2018) (Huang. 2022). and NHANES (2003-
2010) (Hara et al.. 2015) did not identify associations between BLLs and prevalent hypertension.
Additionally, the GuLF study, (n = 957) which cross-sectionally evaluated concurrent BLLs and prevalent
hypertension among those involved in the 2010 Deepwater Horizon oil spill, indicated no associations
(Xu et al.. 2021). However, this study had low (quartile 1: 0.06 (ig/dL, quartile 4: 0.27 (ig/dL) mean
BLLs. These results are consistent with the 2013 Pb ISA, which generally summarized studies reporting
null associations between concurrent BLLs and prevalent hypertension. For example, a study of South
Korean Pb workers indicated no association between BLLs and prevalent hypertension, despite this
population having relatively high BLLs (mean: 31.9 (ig/dL) (Weaver et al.. 2008).
Although most cross-sectional studies did not observe associations between BLLs and prevalent
hypertension, some of these studies did report positive associations. A cross-sectional Brazilian study
evaluated BLLs among adults >40 years and indicated an association between BLLs and prevalent
hypertension (Lopes et al.. 2017a). There were higher odds of hypertension prevalence noted (OR: 1.08
[95% CI: 1.03, 1.14]), for each 1 (ig/dL higher BLLs. Additionally, a KNHANES (2008-2013) study also
indicated a marginal association for each doubling of BLLs for prevalent hypertension (OR: 1.09 [95%
CI: 0.98, 1.22]) (Lee et al.. 2016a). Another more recent analysis using the China National Human
Biomonitoring longitudinal survey evaluated the relationship between concurrent BLLs and several
different definitions of hypertensive status (Ou et al.. 2022). When hypertension was defined according to
the 2010 Chinese Hypertension Guidelines (SBP >140 mmHg, DBP >90 mmHg), there were higher odds
of hypertension associated with BLLs (OR: 2.33 [95% CI: 1.67, 3.24]), when comparing the largest
quartile (>3.2 (ig/dL) with the lowest (<1.5 (ig/dL). Another recent NHANES (1999-2016) analysis (Tsoi
et al.. 2021) indicated higher odds of prevalent hypertension for each doubling of BLLs (OR: 1.09 [95%
CI: 1.04, 1.14]) and when comparing the highest quartile (>2.10 (ig/dL) with the lowest quartile
(<0.89 ng/dL) (OR: 1.21 [95% CI 1.07, 1.36]).
Several studies also evaluated the association between BLLs and prehypertension, a common
precursor to chronic hypertension. Ou et al. (2022) (n = 11,037) considered several prehypertension
definitions. Using the 2010 Chinese Hypertension Guidelines for prehypertension (SBP 120-139 mmHg,
DBP 80-89 mmHg), there were increased odds of prehypertension comparing the highest with the lowest
quartile (OR: 1.56 [95% CI: 1.22, 1.99]). This study also considered the 2017 American College of
Cardiologists (ACC)/American Heart Association (AHA) guidelines for elevated BP (SBP 120-
129 mmHg, DBP <80) and stage 1 hypertension (SBP 130-139 mmHg, DBP 80-89). Using these
definitions, there was a null association between BLLs and elevated BP (OR: 1.18 [95% CI: 0.88, 1.57]),
but a positive association with stage 1 hypertension (OR: 1.75 [95% CI: 1.31, 2.33]). Lee et al. (2016a)
also evaluated prehypertension, which was defined as DBP >80 mmHg or SBP >120 mmHg in a
KNHANES (2008-2013) analysis. This study indicated that for each doubling of BLLs there was an
increased association with prehypertension (OR: 1.09 [95% CI: 0.99, 1.21]). In another KNHANES
(2007-2013) analysis, Lee et al. (2016b) also specifically evaluated prehypertension, which was defined
as DBP between 80-89 mmHg or SBP between 120-139 mmHg and the absence of any current treatment
4-29
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or diagnosis of hypertension. When comparing the highest quartile (2.717 to 24.532 (ig/dL) to the lowest
quartile (0.206 to 1.539 (ig/dL) there was an association between BLLs and prehypertension (OR: 1.30
[95% CI: 1.07, 1.60]).
A recent longitudinal study (Malmo Diet and Cancer Study), within a cohort with high historical
Pb exposure, explored BLLs as they relate to incident hypertension (Gambclunghc et al„ 2016). This
study defined hypertension status as SBP >140 mmHg or DBP >90 mmHg or the use of antihypertensive
medication. At baseline (time = 0) there was a cross-sectional relationship between hypertension and the
highest quartile of BLLs, compared with the lowest three quartiles (OR: 1.3 [95% CI: 1.1, 1.5]).
Participants in this study were followed for approximately 16 years. When analyzed at the follow-up,
there was no association between baseline BLLs and the use of antihypertensive medication (OR: 1.0
[95% CI: 0.8, 1.2]) or high BP at follow-up (OR: 1.0 [0.7, 1.3]).
Another longitudinal analysis evaluated resistant hypertension and both blood and bone Pb levels
among participants of the NAS cohort (Zheutlin et al., 2018). Resistant hypertension was defined as
having either uncontrolled hypertension (SBP >140 or DBP >90 while taking >3 antihypertensive
medications), or controlled hypertension (SBP <140 and DBP <90 while taking >4 antihypertensive
medications). Overall, a 10 (ig/g increase in tibia Pb level was associated with resistant hypertension (RR:
1.12 [95% CI: 1.01, 1.25]) but the association with same increase in patella Pb levels was smaller in
magnitude (RR: 1.04 [95% CI: 0.96, 1.13]). The dose-response relationship between tibia Pb levels and
resistant hypertension risk is relatively linear, with the steepest slope noted in the lower part of the
distribution of tibia Pb concentrations (0 to 20 (ig/g) (Figure 4-13); a flattening of the slope between 20
and 80 |ig/g: and a steepening of the slope for the highest bone Pb concentrations (>80 |ig/g). This dose-
response relationship supports previous findings of a supralinear association between Pb exposures and
Pb-related health outcomes (U.S. EPA, 2013). However, among the same study participants. (Zheutlin et
al., 2018) for a 1 (ig/dL increase in BLLs, the association between BLL and resistant hypertension was
smaller in magnitude (RR: 1.02 [95% CI: 0.97, 1.08]).
4-30
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o -
0 20 40 60 80 100 120
Tibia Lead (pg/g)
HTN = hypertension; RR = relative risk.
Source: Zheutlin et al. (2018).
Figure 4-13 Dose-response curve between tibia Pb levels and resistant
hypertension, Normative Aging Study cohort.
4.3.1.2.1 Effect Measure Modification
Several recent studies also evaluated EMM by a variety of factors when assessing the relationship
between biomarkers of Pb exposure and hypertension outcomes. In a recent NHANES (1999-2006)
analysis, Miao et al. (2020) evaluated EMM by sex for the relationship between BLL and any
hypertension status and uncontrolled hypertension (Figure 4-14, Figure 4-16, Table 4-4). Any
hypertension was defined as SBP >130 mmHg or DBP >80 mmHg or the use of antihypertension
medication, while uncontrolled hypertension was defined as an average SBP >130 mmHg or DBP
>80 mmHg, regardless of antihypertension medication use. When considering continuous BLLs, for each
1 (ig/dL increase in blood Pb there were higher odds of any hypertension among males (OR: 1.037 [95%
CI: 1.015, 1.060]), but less so among females (OR: 1.020 [95% CI: 0.970, 1.074]). However, for each
1 (ig/dL increase in BLLs, there were higher odds of uncontrolled hypertension for both hypertensive
males (OR: 1.157 [95% CI: 1.080, 1.239]) and females (OR: 1.109 [95% CI: 1.020, 1.205]) and a smaller
elevation in the odds of uncontrolled hypertension among all males (OR: 1.062 [95% CI: 1.036, 1.088])
and females (OR: 1.056 [95% CI: 1.011, 1.102]). The dose-response relationship, when considering a
restricted cubic spline for BLLs, indicated a steeper slope up to around 2 (ig/dL, like has been observed
for other Pb exposure and hypertension outcomes (Figure 4-14, See Section 4.3.1.2). This relationship
appears to be more pronounced in males than in females, especially when comparing uncontrolled
4-31
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hypertension to those with controlled hypertension.
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Several other studies also evaluated EMM by sex for the association between biomarkers of Pb
exposure and hypertension. Lee et al. (2016a) evaluated both hypertension and prehypertension using
KNHANES (2008-2013) and observed positive associations between BLLs and hypertension (OR: 1.29
[95% CI: 1.10, 1.51]) and prehypertension (OR: 1.21 [95% CI: 1.06, 1.38]) in females only for each
doubling of BLLs. Similarly, in the cross-sectional analysis of the Malmo Diet and Cancer Study,
Gambclunghc et al. (2016) indicated an elevated effect of prevalent hypertension among females (OR: 1.4
[95% CI: 1.1, 1.7]), compared with males (OR: 1.2 [(0.6, 1.5]). However, Ou et al. (2022) indicated
associations larger in magnitude, but with less precision, among males compared with females
(Figure 4-15) in the China National Human Biomonitoring cohort.
Oi 02 Oi 04 Oi Oi 0-5 O*
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The 2017 ACC/AHA guideline
2010 Chinese HTN guideline
Men
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HTN
The 2017 ACC/AHA guideline 2010 Chinese HTN guideline
Women
ACC = American College of Cardiologists; AHA = American Heart Association; HTN = hypertension; Q = quartile.
Source: Adapted from Qu et al. (2022).
Figure 4-15 Effect measure modification by sex for the association between
quartiles of blood Pb and prevalent hypertension.
Evaluation of EMM by race/ethnicity and other socioeconomic factors was less common in
studies of hypertension, compared with studies examining BP alone. Scinicariello et al. (2011) examined
EMM by both race/ethnicity and sex for the relationship between blood Pb and prevalent hypertension,
using NHANES (1999-2006). Although the overall relationship between hypertension and blood Pb was
null, an association was reported among Black males (OR: 2.69 [95% CI: 1.08, 6.72]) when comparing
those with BLLs at the 10th percentile (<0.6 (ig/dL) to those at the 90th percentile (3.5-10 (ig/dL). Hara et
al. (2015) indicated an overall null association between blood Pb and hypertension, however, an outcome
that persisted even when stratifying by race/ethnicity in NHANES (2003-2010). In addition, the GuLF
study (Xu et al.. 2021) indicated a null association between concurrent blood Pb and hypertension in the
full sample and in analyses stratified by race (Table 4-4).
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Diet has also been considered as an EMM of this association. A recent KNHANES (2013)
analysis evaluated the association between BLLs and hypertension by curry intake (Choi et al.. 2018).
Curcumin, a major component of curry, is known to have anti-inflammatory properties and can act as a
chelating agent for heavy metals, such as Pb. This study defined hypertension as SBP >140 mmHg, DBP
>90 mmHg, or current use of antihypertensive medication. This study indicated a null association
between prevalent hypertension and blood Pb among those who regularly consumed curry (consumed at
least one curry dish/month in the past year) (OR: 1.108 [95% CI: 0.827, 1.485]), for a 1 (ig/dL increase in
BLLs; however, an association was reported in those who did not regularly consume curry (OR: 1.399
[95% CI: 1.054, 1.857]). A previous analysis of the NAS cohort (Elmarsafawv et al.. 2006) evaluated
whether calcium intake affects the relationship between hypertensive status and bone Pb levels. High
calcium intake has been associated with lower BP measurements and it has been hypothesized that
calcium and Pb may interact with one another biologically. Using detailed dietary information to estimate
calcium intake indicated there were moderate associations between either concurrent blood or bone Pb
measurements and prevalent hypertension, but this association did not differ among those with low
calcium intake (<800 mg/d) compared with those with high calcium intake (>800 mg/d).
4.3.1.3 Blood Pressure and Hypertension in Children
The 2013 Pb ISA (U.S. EPA, 2013) indicated that the small body of evidence presented
suggested a relationship between biomarkers of Pb exposure and BP and hypertensive effects in children,
adding to the few studies presented in the 2006 Pb AQCD (U.S. EPA, 2006). Although BP effects are
often more prevalent in adult populations compared with child populations, evidence from earlier studies
suggested BP increases related to Pb biomarkers levels in children and adolescents. In the 2013 Pb ISA
(U.S. EPA, 2013), the strongest evidence of a relationship between Pb biomarkers and increased
childhood BP came from longitudinal studies (Zhang et al., 2012; Gump et al., 2005) and cross-sectional
studies (Gump et al., 2011; Factor-Litvak et al., 1999). More recent data supports the previous findings.
Study-specific details, including Pb biomarker levels, study population characteristics, potential
confounders, and select results from these studies are highlighted in Table 4-5. These details include
standardized results as well as those that could not be standardized based on the information provided in
each paper.
Several recent longitudinal studies highlight associations between increased BP associated with
increased levels of biomarkers of Pb exposure in children. A longitudinal study in Mexico City (n = 457
mother-child pairs) evaluated cord BLLs (GM 4.67 (ig/dL) and maternal bone Pb levels (patella [median:
11.6 |ig/g] and tibia [median 9.3 |ig/g |) 1-month postpartum and subsequently assessed BP in their
offspring (aged 9-15) (Zhang et al„ 2012). The associations between any Pb biomarker and changes in
BP were null, but when evaluating sex as an effect modifier, a 10 |ig/g increase in maternal tibia Pb levels
was associated with increased SBP (1.62 mmHg [95% CI: 0.53, 2.71 mmHg]) and DBP (1.24 mmHg
[95% CI: 0.23, 2.25]) in female children, but not in male children. There was no such association for
4-34
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patella Pb or cord BLLs. Cortical (tibia) bone is reflective of cumulative exposure, whereas trabecular
(patella) bone has a shorter half-life and a higher turnover rate of Pb. Additionally, cord blood is mostly
representative of the BLLs in late-pregnancy and at birth, and not necessarily the BLLs the fetus was
exposed to throughout pregnancy. In addition, a small prospective study (n = 122) among 9.5-year-old
children, described in the 2013 Pb ISA, observed an increase in SBP (12.16 mmHg [95% CI: 2.44,
21.88 mmHg]), but only suggested an increase in DBP (8.54 mmHg [-0.45, 17.35 mmHg])
corresponding with a 1 (ig/dL increase in cord BLLs (GM: 2.56 (ig/dL) (Gump et al.. 2005); even so, null
associations were observed between concurrent blood levels and BP measurements.
In contrast, several recent longitudinal studies including mother-child pairs have been
implemented and generally have yielded null results. Kupsco et al. (2019) assessed blood levels for
several metals, including Pb (mean: 3.7 (ig/dL), during the second trimester of pregnancy and specific
cardiac and metabolic endpoints were evaluated among children, in a small prospective study (n = 548
mother-child pairs). The associations between the natural log of maternal BLLs and children's SBP or
DBP were null. Another larger study (n = 1,511 mother-child pairs) using maternal BLLs evaluated the
association between the erythrocyte fraction (Ery-Pb) in maternal blood and BP among children
(-4.5 years) (Skroder et al.. 2016). The Ery-Pb was assessed at both 14 weeks (median: 73 |ig/kg) and
30 weeks (median: 86 |ig/kg) gestation. Linear regression analyses identified no associations with SBP or
DBP among young children. In addition, another recent study Zhang et al. (2021) of mother-child pairs
(n = 1,194), evaluated BLLs in mothers 24-72 hours after delivery, BP was then subsequently assessed in
children (aged 3-15). Among this cohort, there were null associations between mother's BLL and
children's BP measurements.
Several recent cross-sectional studies that evaluated the association between concurrent BLLs and
BP indicated positive associations. Gump et al. (2011) evaluated BP change as a response to acute stress.
Children aged 9-11 (n = 140) were subjected to a variety of experimental tasks to stimulate the stress
response. Children with higher quartiles of concurrent blood Pb (1.21 to 3.76 (ig/dL) exhibited a greater
change in SBP (7.23 mmHg; 95% CI not reported) compared with children with lower blood Pb (0.14 to
0.68 (ig/dL; 5.3 mmHg) (Table 4-5). An earlier study, included in the 2013 Pb ISA, evaluated children
(n = 281) with higher Pb blood levels (4.1 to 76.4 (ig/dL) from two different towns in Kosovo, when it
was part of Yugoslavia, with high (mean: 37.3 (ig/dL) and low (mean: 8.7 (ig/dL) BLLs. This study
identified a modest association between a 1 (ig/dL increase in concurrent BLLs and SBP (0.05 mmHg
[-0.02, 0.13]) (Factor-Litvak et al.. 1996). Additionally, a recent study from China, evaluated childhood
BP and concurrent child blood Pb (Lu et al.. 2018). Children in this study (n = 590) were recruited from
two regions of similar SES in China, corresponding to an e-waste (high environmental Pb, mean:
7.14 (ig/dL) exposed area (Guiyu) and a reference (low environmental Pb, mean: 3.91 (ig/dL) area
(Haojiang); no association was noted between log-transformed BLLs and either SBP or DBP among these
children.
4-35
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Several recent studies assessing BP and BLLs in children have relied on cross-sectional
nationally representative data sets (NHANES, KNHANES). A large study evaluated seven 2-year
NHANES cycles (1999-2012) among adolescents aged 12-19, with an average BLL of 1.17 (ig/dL (Xuet
al.. 2017). In this cohort, there was no association between BLLs and BP. Another NHANES (2009-
2016) analysis also indicated a null association between BP changes and blood Pb among children aged
8-17 (Desai et al.. 2021). Similarly, a smaller study included three KNHANES cycles (2010-2016)
among adolescents 10-18 years of age with aGM BLL of 1.19 (ig/dL (Ahn et al.. 2018). In this study,
there was no association reported between a doubling of BLLs (log-transformed) and BP or
prehypertension (SBP 120-140 mmHg, DBP 80-90 mmHg). Another NHANES analysis, used five 2-
year NHANES cycles (2007-2016) among children and adolescents aged 8-17. This cohort had a GM of
BLLs ranging between 0.98 (ig/dL and 0.60 (ig/dL from the first (2007-2008) to last (2015-2016)
NHANES cycle evaluated. Similarly, there were no associations between BLLs and BP. However, when
race/ethnicity was considered as an effect modifier, twofold higher BLLs were associated with lower
DBP among Black children (-1.59 mmHg [95% CI: -3.04, -0.16 mmHg]), and higher DBP among white
children (1.38 mmHg [95% CI: 0.40, 2.36 mmHg]) (Yao et al.. 2020).
Several studies also evaluated total peripheral resistance (TPR) and its relationship with
biomarkers of Pb in children. In general, TPR measures the total amount of force circulating blood
imposes on the vasculature in the body and is represented by the ratio between MAP and cardiac output.
Gump et al. (2011) evaluated cardiovascular responses, including sympathetic and parasympathetic
activation, in response to acute stress in children. Children aged 9-11 were subjected to a variety of
experimental tasks to stimulate the stress response. Overall, increased BLL quartiles corresponded to an
increase in TPR. These results support a previous study by Gump et al. (2005). which reported higher Pb
exposures during early childhood. In this study, Gump et al. (2005) indicated that an increase in early
childhood (average age 2.6 years) BLLs was associated with a greater TPR response to acute stress years
later (at 9.5 years of age). Overall, in this cohort, TPR increased with increasing quartiles of BLLs.
Furthermore, BLL was identified as a mediator within this cohort between the relationship between SES
and TPR reactivity. Specifically, Gump et al. (2007) indicated that BLLs may also mediate the association
between SES and the cortical responses to acute stress. Furthermore, when controlling for childhood
BLLs, family income (a measure of SES) was no longer predictive of Cortisol levels.
4.3.2 Toxicological Studies of Blood Pressure and Hypertension
In the 2013 Pb ISA for Pb and previous Pb AQCDs, animal toxicological studies have
consistently demonstrated a relationship between exposure to Pb and increases in BP. Nearly all animal
toxicological studies provided evidence that long-term Pb exposure (>4 weeks), resulting in BLLs less
than 10 (ig/dL, could result in the onset of hypertension (after a latency period) in experimental animals
that persists long after the cessation of Pb exposure (U.S. EPA, 2006). For example, Tsao et al. (2000)
presented evidence for increased systolic and diastolic BP in rats with BLLs somewhat similar to the
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current U.S. adult population (mean 2.15 (ig/dL blood Pb), compared with untreated controls. In addition,
there was a statistically significant, positive trend for increasing BP with increasing BLLs up to 56 (ig/dL,
with the effect leveling off at higher BLLs. There were a number of other studies from previous reviews
demonstrating increases in measures of BP following exposure to Pb (Mohammad et al., 2010; Zhang et
al., 2009; Badavi et al., 2008); Grizzo and Cordellini (2008); (Reza et al., 2008; Bravo et al., 2007; Robles
et al., 2007); Heydari et al. (2006); (Bagchi and Preuss, 2005; Nakhoul et al., 1992). More information on
these studies can be found in Section 4.4.2.2 of the 2013 Pb ISA (U.S. EPA, 2013).
Since the publication of the 2013 Pb ISA, animal toxicological studies with mean blood Pb values
of <30 (ig/dl have further demonstrated a relationship between exposure to Pb and increases in measures
of BP. More specifically, rats with a mean BLL of 13.6 (ig/dl following a 30-day drinking water exposure
had statistically significantly higher SBP (p < 0.05) at 1, 2, 3, and 4 weeks of exposure when compared
with control animals (Fioresi et al., 2014). At the end of the 30-day exposure, these authors also reported
statistically significant increases in SBP, DBP, and MAP (Fioresi et al., 2014). Similarly, Nunes et al.
(2015) reported that rats with an 8.4 |ig/dl mean BLL had statistically significantly higher SBP from 7 to
28 days following a 30-day exposure, relative to control animals. In another multi-day measurement
study, Xu et al. (2015) reported a statistically significant increase (p < 0.05) in SBP and DBP between the
6th and 17th day of a 40-day Pb exposure, but no difference from days 19 to 40. Pb levels in this study
were 19.3 (ig/dl at day 12 and 24.6 |ig/dl on day 40 Xu et al. (2015).
In agreement with the studies that measured BP on multiple occasions throughout exposure, Silva
et al. (2015) reported that rats with a 12.3 (ig/dl BLL had statistically significantly (p < 0.05) higher SBP
following a 30-day exposure relative to control animals. In an additional study, Shvachiy et al. (2018)
exposed rats first through lactation. After weaning, rats were then exposed by drinking water either
continuously until 28 weeks or were given 8 weeks of Pb abstinence and then exposed until 28 weeks.
This study reported a statistically significant increase in DBP and MAP in rats continuously or
intermittently exposed to Pb, as well as a significant increase in SBP in rats continuously exposed to Pb
relative controls. In addition, for both exposure groups, the authors reported a statistically significant
(p < 0.05) decrease in BP regulation as measured by differences in baroflex gain. Notably, a decreased
baroflex response can impair BP recovery (i.e., lowering of BP) following stimulation of chemoreceptors
that increase BP. BLLs in this study were -24 |ig/dl for the constant exposure group and -19 |ig/dl for the
intermittent exposed group (Shvachiy et al„ 2018). Similarly, in a study of rats exposed to Pb through
lactation and weaning, statistically significant increases (p < 0.05) in SBP at timepoints ranging from
PND 22 to PND 100 were reported relative to control animals. Mean BLLs ranged from -11 |ig/dl to
20 (ig/dl in this study (Gaspar and Cordellini, 2014). In agreement with these studies, a pair of analyses
demonstrated a statistically significant increase (p < 0.05) in SBP (but not DBP) relative to control
animals following maternal exposure and then an additional a 1-year drinking water exposure that
resulted in a BLL of <30 (ig/dl (Zhu et al., 2019; Zhu et al., 2018).
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While the above studies all reported some statistically significant increases in BP at one or
multiple timepoints, Wildemann et al. (2015) reported no change relative to control animals for SBP,
DBP, or PP for rats with a 1.7 |ig/dl or 8.6 |ig/dl BLL after 4 weeks of exposure. Moreover, combined
exposure of Pb, mercury, and methylmercury resulted in no change in any of these BP measures relative
to control (Wildemann et al.. 2015).
When considered as a whole, the animal toxicological evidence presented above continues to
demonstrate a clear relationship between Pb exposure and increases in measures of BP. All but one
animal toxicological study evaluated above reported at least some measure of increased BP following Pb
exposure. Additional details on these studies and their designs can be found in Table 4-6.
4.3.2.1 Renin-Angiotensin-Aldosterone System
The renin-angiotensin-aldosterone system (RAAS) plays an important role in the regulation of
BP. For example, angiotensin II (Ang II) stimulates arteriolar vasoconstriction leading to increases in BP.
Angiotensin-converting enzyme (ACE) is involved in the activation of Ang II. In the 2013 Pb ISA, most
studies demonstrated an effect of Pb on RAAS consistent with increases in BP. For example, following
Pb exposure, vascular reactivity to Ang II was found to increase (Robles et al., 2007). Moreover,
exposure to Pb also resulted in increases in kidney and/or serum ACE activity and renal angiotensin II
positive cells (Rodriguez-Iturbe et al., 2005; Sharifi et al., 2004; Carmignani et al„ 1999). In addition, Pb
exposure increased activity and levels of the a-1 subunit protein of Na+/K+ATPase, which plays a major
role in Na+ reabsorption and is regulated by the RAAS (Fiorim et al„ 2011; SimSes et al., 2011). Other
studies demonstrating effects on RAAS can be found in Section 4.4.2.3 of the 2013 Pb ISA (U.S. EPA,
2013).
Since the 2013 Pb ISA, Fioresi et al. (2014) reported a statistically significant increase in NA+ K+
ATPase (p < 0.05) but no change in ACE activity in plasma and cardiac tissue. Thus, there is limited
additional evidence for changes in RAAS following Pb exposure resulting in BLLs <30 (ig/dl. Additional
details for this toxicological study can be found in Table 4-6 of this ISA.
4.3.3 Integrated Summary of Blood Pressure and Hypertension
Several studies presented in the 2013 Pb ISA demonstrated positive associations between BP
measurements and biomarkers of Pb exposure. The current literature continues to support these findings.
Since the 2013 Pb ISA, several nationally representative cross-sectional studies (e.g., NHANES,
KNHANES) have evaluated the association between concurrent blood Pb values and BP measurements or
hypertension status. These studies can contribute substantially to the current evidence base, especially
since there were fewer of nationally representative studies available at the time of the 2013 Pb ISA.
Typically, cross-sectional studies can provide information on the association between concurrent blood
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Pb values and BP measurements or on hypertension status taken at the time of the interview. While cross-
sectional study designs have several limitations, it is important to emphasize the exposure window
reflected in the different Pb biomarkers being considered. Specifically, blood Pb is a better reflection of
more recent exposures and bone Pb is more closely linked with cumulative exposure. However,
longitudinal studies can typically provide information on the relationship between historic Pb biomarker
information and the change in BP since baseline or the development of hypertension. However,
longitudinal studies may be biased if there is a large loss of follow-up. Both study types are valuable in
discerning the associations between biomarkers of Pb exposure, BP, and hypertension. Overall, recent
cross-sectional studies provided consistent evidence that higher concurrent BLLs are associated with
higher SBP and DBP within adult populations. Evidence for higher PP and MAP were less consistent but
these endpoints were examined in fewer studies. Associations between concurrent BLLs and BP among
children were inconsistent, and mostly suggested a null association. Yet, a series of studies evaluating
TPR among children indicated an association with increasing blood Pb values. In addition, studies
evaluating a concurrent BLL and BP at a particular threshold (i.e., SBP >130 mmHg), mostly indicated
null results.
Longitudinal studies less commonly evaluated changes in BP measurements (in mmHg) but were
more likely to evaluate the development of clinical hypertension or prehypertension over a prolonged
period of time. Most longitudinal studies evaluating incident hypertension or prehypertension and a
marker of cumulative Pb exposure (measured in bone) indicated positive associations. In contrast,
associations with incident hypertension or prehypertension were mostly null when using blood Pb
measurements. Of the few longitudinal studies that evaluated BP changes, the results were mostly null,
with a few indicating associations between baseline blood Pb measurements and changes in BP
measurements overtime.
Animal toxicological studies continue to support the epidemiologic evidence. Recent animal
toxicological studies reaffirm the clear association between Pb exposure in animals and increases in BP,
presented in the 2013 Pb ISA. Current studies specifically were restricted to only include lower BLLs
(<30 (ig/dL), and the majority of relevant studies indicated a persistent relationship between BLLs,
whether it be related to continuous or intermittent exposures, and increases in BP. The evidence
supporting changes in RAAS following Pb exposure is less consistent.
Several recent epidemiologic studies also evaluated EMM by race/ethnicity, sex, age genetic
polymorphisms, among others. Taken together the evidence suggests that in addition to having higher
blood Pb measurements, associations between blood Pb and BP are larger among non-Hispanic Black
populations when compared with Hispanic or non-Hispanic white populations. When considered alone,
there were mixed conclusions as to whether there were any differences in the association between Pb
biomarkers and BP or hypertension by sex. However, when combined with race, Black males clearly
demonstrated increased risk of Pb-associated BP changes, when compared with other sex/race groups.
These results were consistent across several analyses. Taken together, the most recent evidence supports
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the conclusions of the previous ISA, indicating an association between biomarkers of Pb exposure and
changes in either BP or hypertension status, with evidence that certain populations may be at increased
risk.
4.4 Ischemic Heart Disease and Associated Cardiovascular
Effects
IHD, also known as CHD or CAD, is a chronic condition characterized by atherosclerosis and
reduced blood flow to the heart. The majority of IHD is caused by atherosclerosis (Section 4.8), which
can lead to the blockage of the coronary arteries and restriction of blood flow to the heart muscle. An MI
or heart attack is an acute event that occurs when heart tissue death occurs secondary to prolonged
ischemia. Several studies within this section evaluate IHD as a composite measure mostly defined as the
presence of MI, angina pectoris, or CHD death, whereas other studies evaluate composite IHD- risk
scores using cross-sectional data. There were no animal toxicological studies examining indicators of IHD
at BLLs <30 (ig/dL published since the 2013 Pb ISA.
4.4.1 Epidemiologic Studies of Ischemic Heart Disease
The 2006 Pb AQCD (U.S. EPA, 2006) indicated an association between Pb biomarker levels and
MI (Gustavsson et al., 2001). The 2013 Pb ISA (U.S. EPA, 2013) further contributed to this small
amount of evidence with the inclusion of a study among the NAS cohort. This longitudinal study among
(mostly white) men indicated an increased incidence of IHD associated with bone (both tibia and patella)
Pb levels (Jain et al„ 2007).
Several recent studies have been published since the 2013 Pb ISA that specifically evaluate the
association between biomarkers of Pb exposure and measures of IHD, CHD, or CAD. Study-specific
details, including biomarker Pb levels, study population characteristics, confounders, and select results
from these studies are highlighted in Table 4-7. These details include standardized results as well as those
that could not be standardized based on the information provided in each paper.
A recent study evaluated whether the relationship between CHD and bone Pb levels is modified
by certain genetic polymorphisms (Ding et al., 2016). It is thought that certain genetic factors may
predispose an individual to increased Pb toxicity. Using the NAS cohort, several genes and encoding
proteins including ALAD, HFE, heme oxygenase-1 (HMOX1), VDR, apolipoprotein E (APOE),
glutathione S-transferases, and the RAAS, were evaluated as effect measure modifiers of the relationship
between bone Pb measurements and incident CHD. All these different genes and encoding proteins
appear to play a role in influencing Pb uptake and or retention or may alter Pb toxicity. Overall, 22
different SNPs corresponding to these Pb-related genes were studied separately and in combination in a
genetic risk score (GRS). Two GRSs were constructed; the first (GRS 1) summed all 22 SNPs, whereas
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the second (GRS 2) only included the nine SNPs found to significantly modify the association between
patella Pb levels and incident CHD within this study. Overall, without considering any genetic
polymorphisms, the association between a twofold increase in patella Pb levels and CHD incidence was
positive (HR: 1.36 [95% CI: 1.15, 1.61]). Several genetic polymorphisms appeared to further modify this
relationship. Specifically, positive associations were observed for individuals with at least one minor
allele in VDR (rsl544410 (Bsml) HR: 1.65 [95% CI: 1.31, 2.08]); rs731236 (Taql) HR: 1.61 [95% CI:
1.29, 2.02]); rsl073581 (Fokl) HR: 1.47 [95% CI: 1.17, 1.83]); rs757343 (Tru91) HR: 1.48 [95% CI:
1.18, 1.85]) and HMOX1 (rs2071749) HR: 1.51 [95% CI: 1.22, 1.86]), whereas individuals without any
minor alleles had null associations. However, positive associations were observed among individuals
without any minor alleles in HMOX1 (rs2071746 HR: 1.51 [95% CI: 1.07, 2.13]; rs5995098 HR: 1.62
[95% CI: 1.23, 2.14]), APOE (rs429358 HR: 1.43 (95% CI: 1.17, 1.75]) and angiotensinogen (AGT;
rs699 HR: 2.17 [95% CI: 1.51, 3.12]; rs5046 HR: 1.57 [95% CI: 1.27, 1.94]). When considered in
combination, both GRS values identified significant EMM for the association between a twofold increase
in patella Pb and risk of incident CHD (GRS 1 HR: 2.27 [95% CI: 1.50, 3.42] and GRS 2 HR: 2.77 [(95%
CI: 1.78,4.31]).
Another study of the NAS cohort measured if incident CAD and bone Pb measurements were
modified by diet (Ding et al.. 2019). Evidence suggests that a diet deficient in essential metals (zinc,
calcium, selenium, iron) can augment Pb absorption and retention in the body, while certain vitamins (C,
E, and Be) may function as antioxidants against Pb toxicity. Specifically, vitamins E and C can act by
inhibiting lipid peroxidation by neutralizing Pb-related reactive oxygen species (ROS) by rapid electron
transfer, while vitamin Be can act by reducing Pb-related increases in homocysteine. Additionally,
vitamins Bi and Be are composed of ring structures containing nitrogen, which may mediate interactions
with Pb. This study collected detailed dietary information from each NAS member and classified diets
high in fruit, legumes, whole grains, tomatoes, seafood, poultry, cruciferous vegetables, dark-yellow
vegetables, leafy vegetables, and other vegetables as "prudent" diets. Alternatively, diets with a high
intake of processed meat, red meat, refined grains, butter, high-fat dairy products, eggs, and fries, was
considered a "Western" diet. The diet types were considered separately and were not mutually exclusive.
For example, a low prudent diet was not equivalent to a high Western diet, and there could be some
overlap in diet type between participants. Overall, results indicated that for each doubling of bone Pb
levels there was a higher risk of CAD associated with both tibia (HR: 1.25 [95% CI: 1.06, 1.48]) and
patella (HR: 1.30 [95% CI: 1.09, 1.56]). However, low prudent diet modified this association with patella
Pb levels. Those with a low prudent diet (HR: 1.64 [95% CI: 1.27, 2.11]) had a higher association
between patella Pb levels and CAD risk compared with those with a high prudent diet (HR: 1.07 [95% CI:
0.86, 1.34]). A Western diet did not appear to modify the results.
In a Canadian prospective cohort of patients on hemodialysis, incident cardiovascular events
during the 2-year follow-up period were evaluated (Tonelli et al.. 2018). Cardiovascular events were
defined as acute MI, percutaneous coronary angioplasty, coronary artery bypass grafting, heart failure,
and stroke or transient ischemic attack. Patients in this cohort (n = 1,278) had relatively low BLLs (1st
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decile: 0.06 (ig/dL, 10th decile 1.74 (.ig/dL). and there was no observed relationship between BLLs and
cardiovascular events when comparing the highest with the lowest decile (results not shown).
Several recent cross-sectional analyses have assessed 10-year CHD risks in association with
biomarkers of Pb exposure (Nguyen et al.. 2021; Park and Han. 2021; Choi et al.. 2020; Cho et al.. 2016).
Cho et al. (2016) calculated the Framingham risk score (FRS) to predict the 10-year risk of CHD in
asymptomatic patients associated with BLLs among Korean men and women taking part in KNHANES
IV and V (2008-2010). The FRS incorporates various CHD risk factors including age, gender, SBP, total
cholesterol, and high-density lipoprotein cholesterol (HDL-C). This study indicated that for each
increasing BLL quartile, there were statistically significant increased odds of an elevated FRS, compared
with the lowest quartile among men. Specifically, there was a positive effect (OR: 3.13 [95% CI: 2.09,
4.69]) for the highest quartile of BLLs (3.519-26.507 (ig/dL) compared with the lowest quartile of BLLs
(0.711-2.129 (ig/dL) among males. This effect was not observed among females (OR: 0.88 [95% CI:
0.26, 2.97]). Park and Han (2021) also calculated a CVD risk score based of the FRS from 2008. Again,
using data obtained from KNHANES, this study calculated the effect of a log increase in BLLs associated
with a 10%-20% increase in FRS. Park and Han (2021) indicated that a one-unit increase in log-
transformed blood Pb was associated with an odds ratio of 2.4 (95% CI: 1.89, 3.18) of having an FRS
increase between 10%-20% in men. However, this association was not observed among females (OR:
1.05 [95% CI: 0.68, 1.63]). Similarly, a >20% increase in the FRS score was associated with an odds ratio
of 2.85 (95% CI: 2.02, 4.01) among males, but not among females (OR: 0.71 [0.19, 2.66]). Another
assessment of KNHANES indicated that a doubling of BLLs was associated with an 0.10% (0.02,
0.21%]) increase in 10-year CVD risk (Nguyen et al.. 2021) (Table 4-7)
(Choi et al.. 2020) used KNHANES to evaluate associations between BLLs and the 10-year
atherosclerotic cardiovascular disease (ASCVD) risk score. The ASCVD risk score was calculated first
based of the ACC and AHA guideline on the assessment of CVD risk. This formula incorporates factors
such as age, total cholesterol, HDL-C, hypertension treatment, smoking status, and diabetes. For this
analysis, the risk score was scaled to be more relevant to the Korean population, as the risk score was
created based on mostly non-Hispanic white and non-Hispanic Black populations in the United States.
This study also noted a higher ASCVD risk of 0.117 (95% CI: 0.005, 0.229) among men when comparing
the highest with the lowest quartiles (distribution information not reported), but not among women
(0.072, [95% CI: -0.004, 0.148]). When EMM was considered for urban versus rural locations, there was
a higher ASCVD risk score effect estimate among men living in urban areas (0.133 [95% CI: 0.011,
0.254]) and among women living in rural communities (0.212 [95% CI: 0.045, 0.379]).
A recent cross-sectional study evaluated older, diabetic patients in China (Wan et al.. 2021). This
study (n = 4,324) evaluated BLLs and prevalent CVD. In this context, CVD was defined as a composite
measure including a history of CHD, MI, or stroke. When comparing the highest quartile of BLLs
(>3.7 (ig/dL) with the lowest quartile of BLLs (<1.8 (.ig/dL). there were higher odds (OR: 1.44 [95% CI:
1.17, 1.76]) of CVD within this population at elevated BLLs (Figure 4-16). Additionally, another recent
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study (n = 175) evaluated a collection of emerging predictive CVD biomarkers including asymmetric
dimethylarginine (ADMA), adipocyte fatty acid-binding protein (FABP4, also known as aP2 and
AFABP), adiponectin, and chemerin (Ochoa-Martincz et al.. 2018). When comparing the highest tertile
(T3: >9.1 (ig/dL) with the lowest tertile (Tl: <3.5 (.ig/dL), there was a positive association with ADMA
(0.75 (miol/L [95% CI: 0.15, 1.85 ^mol/L]) and FABP4 (27.5 ng/mL [95% CI: 10.0, 34.5 ng/mL]). Other
biomarkers evaluated had null associations with BLLs.
CCA plaque
P for trend < 0.001
One In BLL SD increment
BLL Quartile 4
BLL Quartile 3
BLL Quartile 2
BLL Quartile 1
0.5 1.0 1.5
Odds ratio (95%CI)
2.0
Left CCA diameter
Pfor trend < 0.384
One In BLL SD increment
BLL Quartile 4
BLL Quartile 3
BLL Quartile 2
BLL Quartile 1
-0.2 -0.1 0 0.1 02
regression coefficients (95%CI)
CVD
P for trend < 0.001
One In BLL SD Increment
BLL Quartile 4
BLL Quartile 3
BLL Quartile 2
BLL Quartile 1
0.5 1.0 1.5
Odds ratio (95%CI)
2.0
Right CCA diameter
P for trend = 0.777
One In BLL 3D increment
BLL Quartile 4
BLL Quartile 3 i-
BLL Quartile 2
BLL Quartile 1
-0.2 -0.1 0 0.1 0.2
regression coefficients (95%CI)
BLL = blood lead level; CCA = common carotid artery; CI = confidence interval; CVD = cardiovascular disease; SD standard
deviation.
Source: Wan et al. (2021).
Figure 4-16 Relationship between blood Pb levels and common carotid artery
plaques, common carotid artery diameter, and cardiovascular
disease among diabetic patients.
A recent meta-analysis evaluating blood metals (including blood Pb) evaluated the aggregate
association between BLLs and CHD risk (Chowdhurv et al.. 2018). For this study, CHD was defined as
non-fatal MI, angina, coronary revascularization (i.e., percutaneous transluminal coronary angioplasty or
coronary artery bypass surgery) or CHD mortality. It included studies with cohort, case-control, or
nested-case-control study designs. In this analysis, a total of eight studies were identified including those
that evaluated CHD mortality, those that were previously included in the 2006 Pb AQCD (U.S. EPA.
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2006), and occupational studies. Even though none of the studies presented in this meta-analysis of CHD
and BLLs were included in this section (cardiovascular mortality discussed in Section 4.10), the overall
results further support an association between biomarkers of Pb exposure and IHD (Figure 4-17).
Measurement
source
No of
participants
Lead
No of
events
SOF
Blood
533
54
Glostrup Population Studies
Blood
1050
54
Zutphen study
Blood
146
64
VA-NAS
Blood
1235
185
BRHS
Blood
7379
382
NHANES II
Blood
4190
424
ABLES
Blood
58368
692
McElvenny (2015)
Blood
9122
941
NHANES III
Blood
18602
985
NHANES III
Blood
9757
1189
Relative risk
(95% CI)
Subtotal: P=0.001, I =67.6%
ABLES = Adult Blood Lead Epidemiology and Surveillance; BRHS = British Regional Heart Study; CI =
NHANES = National Health and Nutrition Examination Survey; SOF = Study of Osteoporotic Fractures;
Normative Aging Study.
Source: Adapted from Chowdhurv et al. (2018).
Relative risk
(95% CI)
2.23 (0.99 to 4.99)
1.11 (0.61 to 2.00)
1.05 (0.53 to 2.10)
0.69 (0.32 to 1.47)
1.21 (0.82 to 1.78)
1.25 (1.00 to 1.57)
1.64 (1.21 to 2.22)
4.09 (2.48 to 6.74)
1.24 (1.03 to 1.50)
1.47 (1.14 to 1.89)
1.43 (1.16 to 1.76)
confidence interval;
VA-NAS = Veterans Affairs
Figure 4-17 Meta-analysis of the association between biomarkers of Pb
exposure and coronary heart disease.
4.4.2 Summary of Ischemic Heart Disease
Limited evidence was presented in the 2013 Pb ISA (U.S. EPA, 2013) indicating an association
between biomarkers of Pb exposure and incident IHD. Although this effect was strong across both blood
and bone (patella) Pb measurements, there were not enough published studies at the time to fully evaluate
the association.
Several recent epidemiologic studies have been published further supporting this association.
Studies using the NAS cohort of elderly (mostly white) men indicated a positive association between
patella Pb levels and incident IHD (Ding et al„ 2019; Ding et al„ 2016). These studies had extensive
follow-up periods (~20 years), with patella Pb levels ranging between 29.2 and 32.2 ug/g. Additionally, a
series of 10-year CVD risk evaluations ( Nguyen et al., 2021; Park and Han. 2021; Choi et aL 2020; Cho
et al.. 2016) were conducted using KNHANES data. These studies used cross-sectional data to create a
score that could be predictive of future CVD risk, and all indicated higher 10-year CVD risk with
increasing BLLs. BLLs in these studies generally averaged <3 (ig/dL.
While many of these studies evaluated the overall associations between biomarkers of Pb
exposure and IHD or other similar outcomes, many evaluated EMM by sex, diet, and other distinguishing
characteristics such as genetic polymorphisms. Overall, males (Park and Han, 2021; Choi et al... 202.0;
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Cho et al.. 2016) tended to have larger Pb-associated IHD risks than females and certain genetic
polymorphisms (Ding et al.. 2016) modified the relationship between bone Pb levels and incident IHD.
Furthermore, associations between bone Pb levels and incident IHD were larger for people with diets low
in fruit, whole grains, and vegetables (Ding et al.. 2019).
4.5 Heart Failure and Impaired Cardiac Function
Heart failure refers to a set of conditions in which the heart's pumping action is weakened. With
congestive heart failure (CHF), the flow of blood from the heart slows and fails to meet the oxygen
demands of the body, and the returning blood can back up and cause swelling or edema in the lungs or
other tissues (typically in the legs and ankles). Right-sided heart failure is typically a consequence of left-
sided heart failure but can also result from damage to the pulmonary vasculature, which can result in
increased right ventricular (RV) mass, reduced flow to the left ventricle, and reduced left ventricular (LV)
mass. In chronic heart failure, the heart typically enlarges and develops more muscle mass. The 2006 Pb
AQCD (U.S. EPA, 2006) presented limited epidemiologic evidence on the association between
biomarkers of Pb exposure and cardiac function. Little evidence was added in the 2013 Pb ISA. Since
then, the evidence has expanded modestly, with recent epidemiologic and toxicological studies providing
support for an effect between biomarkers of Pb exposure and cardiac function.
4.5.1 Epidemiologic Studies of Impaired Cardiac Function
The 2006 Pb AQCD presented a cross-sectional study indicating an association between Pb
biomarker levels and LV hypertrophy (Schwartz. 1991). More recent studies indicate an association
between Pb biomarkers and cardiac function. Study-specific details, including biomarker Pb levels, study
population characteristics, potential confounders, and select results from these studies are highlighted in
Table 4-8. These details include standardized results as well as those that could not be standardized based
on the information provided in each paper.
A recent small (n = 179) prospective study (Yang et al.. 2017) of a Flemish population evaluated
potential toxic effects of Pb on the myocardium by assessing the association between blood Pb and LV
function. Doppler imaging of transmittal blood flow was used to assess systolic and diastolic LV
function. In this study, there was evidence of decreased LV systolic function for each doubling of blood
Pb. Specifically, there were decreases in global longitudinal strain (GLS) by 0.497% (95% CI: -0.957,
-0.038%), regional longitudinal strain (RLS) by 0.784% (-1.482, -0.087%), regional radial strain (RRS)
by 2.316% (-4.748, -0.115%), and regional longitudinal strain rate by 0.071s 1 (95% CI: -0.124,
-0.019s '). There was no association between BLLs and diastolic LV function. A cross-sectional study
(n = 993) among a Swedish population evaluated LV measurements using two-dimensional
echocardiography measuring septal thickness, posterior wall thickness, LV diameter in end diastole, and
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LV diameter in end systole (Lind et al.. 2012). For natural log increases in serum Pb, there was lower LV
mass index (LVMI) (|3: -0.73 [95% CI: -2.20, 0.74]) and higher relative wall thickness (RWT) (|3: 0.011
[95% CI: -0.001, 0.022]), but neither were statistically significant.
4.5.1.1 Impaired Cardiac Function in Children
The 2013 Pb ISA (U.S. EPA, 2013indicated that the small body of available evidence suggested
a relationship between biomarkers of Pb exposure and cardiac function in children, adding to the few
studies presented in the 2006 Pb AQCD (U.S. EPA, 2006). Specifically, Gump et al. (2011) evaluated
cardiovascular responses, including sympathetic and parasympathetic activation, to acute stress in
children. Children aged 9-11 were subjected to a variety of experimental tasks to stimulate the stress
response. Cardiovascular measurements, including cardiac output and stroke volume were assessed at
baseline and following each task. In general, increasing quartiles (Ql: 0.14-0.68 (ig/dL, Q4: 1.21-
3.76 (ig/dL) of BLLs corresponded to decreases in stroke volume and cardiac output, compared with
baseline. These results support a previous study by Gump et al. (2005), which had higher Pb exposures
during early childhood.
A recent cross-sectional study provided further evidence of an association between more sensitive
cardiac outcomes (Chen et al., 2021). This study evaluated Pb's potential effect on structural function and
inflammation related to LV function in children. Children were recruited from two different primary
areas, including an e-waste exposed area (Guiyu) and a reference area (Haojiang). Several different LV
measurements were obtained. A 1-unit increase in BLL was associated with smaller (natural log)
interventricular septum (IVS) measurements (|3: -0.004 (95% CI: -0.007, -0.001). Other natural log
echocardiogram measurements indicated null associations (LV posterior wall |3: -0.001 [95% CI: -0.003,
0.001]); ejection fraction |3: -0.001 (95% CI: -0.002, 0.001) with a unit increase in BLL.
4.5.2 Toxicological Studies of Impaired Cardiac Function
The previous ISA did not include any animal toxicological studies examining impaired cardiac
function. However, animal toxicological studies published since the last review have looked at the
potential for Pb exposure to alter cardiac function. Wildemann et al. (2015) reported no evidence for an
effect of Pb exposure on stroke volume or cardiac output in rats. Moreover, combined exposure to Pb,
mercury, and methylmercury resulted in no change in these measures relative to controls. In contrast,
Fioresi et al. (2014) reported a statistically significant increase in some measures of cardiac contractility
in rats. More specifically, they found a statistically significant increase in left ventricular systolic pressure
(LVSP) and LV dP/dt, (change in pressure/change in time; p < 0.05), but not right ventricular systolic
pressure (RVSP) or RV dP/dt following a 30-day exposure to Pb (13.6 (ig/dl mean BLL). There were also
no changes reported in left or right ventricular diastolic pressure (LVDP, RVDP) (Fioresi et al.. 2014).
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Additional studies examining the potential for impaired cardiac function were done in isolated
LV papillary muscle. Silva et al. (2015) reported no significant difference in force generation between
muscle isolated from control or Pb-treated (15-day exposure, 12.3 |ig/dl BLL) rats following pulse
stimulation. However, the time to peak tension and 90% relaxation was statistically significantly
(p < 0.05) shorter in LV papillary muscle derived from Pb-treated animals relative to muscle from control
animals. Moreover, inotropic contractile force was statistically significantly decreased in muscle from Pb-
treated animals following treatment with calcium chloride, but not isoproterenol (Silva et al.. 2015) and
Pb exposure significantly lowered tetanic (sustained) peak and plateau force. In a similar analysis in LV
papillary muscle, Fioresi et al. (2014) reported that following a 30-day exposure to Pb resulting in a mean
13.6 (ig/dl BLL, there were not significant differences in isometric contraction force, time to peak
contraction or relaxation rates. However, in contrast to Silva et al. (2015). following rest and calcium
treatment, there was a statistically significant increase in contractile force in muscle from Pb-treated
animals (Fioresi et al.. 2014). When considered as a whole, the animal toxicological evidence for changes
in cardiac function is limited and, in some cases, results across studies are conflicting. Additional details
for the toxicological studies discussed in this section can be found in Table 4-9 of this ISA.
4.5.3 Integrated Summary of Impaired Cardiac Function
Limited evidence was presented in the 2013 Pb ISA (U.S. EPA, 2013) indicating an association
between biomarkers of Pb exposure and indicators of cardiac function. The recent epidemiologic evidence
suggests that the potential effect of Pb exposure on cardiac function may be more likely among children
and the elderly. An analysis of participants >70 years of age indicated positive associations between
markers of LV function and blood Pb, with relatively low mean BLLs (<2 (ig/dL) (Lind et al., 2012).
Associations with these same outcomes were null in a cohort of middle-aged participants (mean age
-39 years), although there was evidence of an association with markers of LV structure within this cohort
(Yang et al„ 2017). A study among children indicated a relationship between smaller IVS measurements
with increased BLLs (Chen et al., 2021). Results of available animal studies examining cardiac function
have been inconsistent, and conflicting results were reported in studies examining contractile force in
isolated papillary muscle following calcium treatment (Silva et al„ 2015; Fioresi et al., 2014).
A small number of studies presented associations between decreased stroke volume with
increasing BLLs (Gump et al„ 2011; Gump et al„ 2005), but results were less consistent when
considering cardiac output. In animal toxicological studies there was no evidence of an effect of Pb
exposure on stroke volume or cardiac output, but limited evidence for an effect on measures of cardiac
contractility. Taken together, there is limited evidence to support a relationship between biomarkers of Pb
exposure and cardiac function.
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4.6 Endothelial Dysfunction
Endothelial dysfunction is the physiological impairment of the inner lining of blood vessels that is
characterized by an imbalance between vasodilators such as nitric oxide and vasoconstrictors such as
endothelin-1 (ET-1). High BP is often the result of an imbalance of these factors that leads to greater
vasoconstriction.
4.6.1 Toxicological Studies of Endothelial Dysfunction
In the 2013 Pb ISA, animal toxicological studies provided mixed evidence for Pb exposure
having an effect on vascular relaxation and constriction. For example, although Pb exposure decreased
acetylcholine (ACh)-induced vasodilation in isolated rat tail arteries (Silveira et al., 2010; Zhang et al.,
2007), Skoczynska and Stojek (2005) reported that Pb exposure enhanced vasodilation by ACh in rat
mesenteric arteries. Moreover, in aortic rings of perinatally exposed rats, there was no change observed in
the relaxation response to ACh (Fiorim et al., 2011; Rizzi et al., 2009; Grizzo and Cordellini, 2008). More
information on these and other studies examining vascular reactivity from previous reviews can be found
in Section 4.4.2.3 of the 2013 Pb ISA (U.S. EPA. 2013).
Since the publication of the 2013 Pb ISA, additional toxicological studies of vascular function
have been published in animals with BLLs <30 (ig/dl. In a study of young rats exposed to Pb through
lactation, Pb exposure (BLL of ~11 |ig/dl to 20 |ig/dl) resulted in a statistically significant (p < 0.05)
increase in the maximum contractile response to the vasoconstrictor noradrenaline in intact rat aortas at
days 52, 70, and 100 (but not at day 23) relative to control animals (Gaspar and Cordellini, 2014). In
denuded aortas (i.e., aortas with no endothelium), the contractile response to noradrenaline increased
comparably from both control and Pb-treated animals, thereby suggesting that the difference in the
contractile response in intact aortas was the result of Pb's effect on the endothelium (Gaspar and
Cordellini, 2014). However, in an additional study using adult rats, there was no difference in intact rat
aortic segments from control or Pb-exposed rats when treated with the vasodilators ACh or sodium
nitroprusside, and a statistically significant (p < 0.05) decrease in the contractile response following
exposure to the vasoconstrictor phenylephrine, but not potassium chloride (Nunes et al., 2015). The BLL
in this study was 8.4 (ig/dl and when the endothelium was mechanically removed, phenylephrine-induced
contractility increased in both groups but to a greater extent in aortic segments from Pb-treated rats. Using
a number of chemical inhibitors, the authors suggest that the decrease in contractility in response to
phenylephrine (in intact aortic segments) was not due to a Pb effect on the vasodilator NO, but rather to
increasing levels of hydrogen peroxide, which can also have vasodilatory effects. That is, incubation with
catalase increased the constriction response to phenylephrine in aortic segments from Pb-treated rats but
not control rats. The authors go on to show that differences in hydrogen peroxide activity between aortic
segments from Pb-treated and control rats is potentially due to Pb increasing the levels of the hydrogen
peroxide generating enzyme superoxide dismutase (SOD) (Nunes et al„ 2015).
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4.6.2
Summary of Endothelial Dysfunction
Taken together, the limited toxicological evidence presented above suggests that Pb exposure
may result in changes in endothelial function. However, the direction of this response varies in that Pb
exposure can either increase or decrease the response to vasodilators/vasoconstrictors. These studies also
suggest that Pb's effects on the endothelium are complicated and differ depending on age, treatment
(e.g., vasodilators testing endothelium-dependent versus endothelium-independent mechanisms), and/or
type of endothelial cells tested. Additional details for the toxicological studies discussed in this section
can be found in Table 4-10 of this ISA.
4.7 Cardiac Electrophysiology and Arrythmia
Electrical activity in the heart is crucial for regulating the heartbeat and is typically measured
using surface electrocardiography (ECG). ECGs measure electrical activity in the heart that is due to
depolarization and repolarization of the atria and ventricles. Changes in electrical activity can lead to
changes in cardiac depolarization, repolarization, and development of arrythmia (Section 4.7.1) and
changes in heart rate and HRV (Section 4.7.2)
4.7.1 Cardiac Depolarization, Repolarization, and Arrythmia
Experimental and epidemiologic studies typically use surface ECGs to measure electrical activity
in the heart resulting from depolarization and repolarization of the atria and ventricles. The P-wave of the
ECG corresponds to atrial depolarization, the QRS complex represents ventricular depolarization, and the
T-wave represents ventricular repolarization. The ventricles account for the largest proportion of heart
mass overall and thus are the primary determinants of the electrical activity recorded in the ECG.
Therefore, ECG changes indicating abnormal electrical activity in the ventricles are of greatest concern.
Endpoints denoting ventricular electrical activity include QTc interval, transmural dispersion duration,
and T-wave shape. Changes in QT and ST, as well as changes in T-wave shape, duration, or amplitude,
may indicate abnormal impulse propagation in the ventricles.
Cardiac arrhythmias can vary in severity from the benign to the potentially lethal, such as in
cardiac arrest when an electrical disturbance disrupts the heart's pumping action causing loss of heart
function. Atrial fibrillation (AF) is the most common type of arrhythmia. Clinical and subclinical forms of
AF are associated with reduced functional status and quality of life, as well as downstream consequences
such as ischemic stroke (Prvstowskv et al.. 1996; Anonymous. 1994) and CHF (Roy et al.. 2009).
contributing to both cardiovascular disease and all-cause mortality (Kannel et al.. 1983). Ventricular
fibrillation is a well-known cause of sudden cardiac death and is commonly associated with MI, heart
failure, cardiomyopathy, and other forms of structural (e.g., valvular) heart disease. Pathophysiologic
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mechanisms underlying arrhythmia include electrolyte abnormalities, modulation of the autonomic
nervous system (ANS), membrane channels, gap junctions, oxidant stress, myocardial stretch, and
ischemia. Ventricular conduction and repolarization abnormalities such as QRS complex and QT interval
prolongation, as well as LV hypertrophy and clinical antecedents including hypertension, are also
associated with cardiac arrest (Rautahariu et al.. 1994).
4.7.1.1 Epidemiologic Studies of Cardiac Depolarization, Repolarization, and
Arrythmia
Numerous epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA, 2013) strengthened
the evidence presented in the 2006 Pb AQCD (U.S. EPA, 2006) that described an association between
biomarkers of Pb exposure and changes in ECG measures. Current studies continue to support prior
analyses. Study-specific details, including blood and bone Pb levels, study population characteristics,
potential confounders, and select results from these studies, are highlighted in Table 4-11. These study
details include standardized results as well as results that could not be standardized based on the
information provided in each paper.
Previous ISAs described analyses evaluating the association between Pb biomarkers and
electrophysiologic outcomes using the NAS cohort, of mostly white men. For example, Cheng et al.
(1998) described an association between bone Pb and corrected QT interval (QTc) among men >65 years,
and Eum et al. (2011) prospectively evaluated ECG findings and bone Pb levels within the NAS cohort.
Eum et al. (2011) reported an association between tibia Pb levels and increases in QTc interval (7.94
msec [95% CI: 1.42, 14.45]) and QRSc duration (5.94 msec [95% CI: 1.66, 10.22]) when comparing the
highest tertile of bone Pb levels with the lowest. Additionally, a cross-sectional analysis of elderly NAS
men provided evidence of EMM of certain genetic polymorphisms in genes affecting iron (Fe)
metabolism (HFE C282Y and HMOX1 L variants) on the relationship between biomarkers of Pb exposure
and prolonged QT interval (Park et al„ 2009).
A recent study supports these previous findings in a more diverse population (NHANES),
compared with the NAS cohort of mostly white men. Jing et al. (2019) used NHANES III (1988-1994) to
evaluate the relationship between log-transformed BLLs and the QRS-T angle. The QRS-T angle can
quantify the relationship between ventricular depolarization (QRS-axis) and repolarization (T-axis) and is
a predictor of ventricular arrythmia. The QRS-T angle was measured using a standard 12-lead ECG, and
sex-specific tertiles of QRS-T angle were created. This study indicated that higher BLLs were associated
with a greater QRS-T angle (third tertile versus first tertile) among men (OR: 1.35 [95% CI: 1.05, 1.74]),
but not among women (OR: 1.05 [95% CI: 0.82, 1.36]).
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4.7.1.2 Toxicological Studies of Cardiac Depolarization, Repolarization, and
Arrythmia
The 2013 Pb ISA evaluated an animal toxicological study demonstrating that exposure to Pb
resulted in increased incidence of arrhythmia and atrioventricular conduction block (i.e., disruption of
electrical signals from the atria to the ventricles) after 12 weeks of Pb exposure (Reza et al.. 2008). This
study also reported a prolonged ST interval, without alteration in QRS duration. Since the last review,
Wildemann et al. (2015) reported no change relative to control animals for PR, QRS, or QT for rats with a
1.7 (ig/dl or 8.6 |ig/dl BLL. A combined exposure of Pb, mercury, and methylmercury resulted, however,
in significant increases in the QRS and QT intervals. Taken together, there is little animal evidence for an
effect of Pb exposure alone on cardiac depolarization and/or repolarization.
4.7.2 Heart Rate and Heart Rate Variability
Heart rate is a key indicator of autonomic function. It is modulated at the sinoatrial node of the
heart by both parasympathetic and sympathetic branches of the ANS and represents the number of times
the heart beats in a given time frame (e.g., per minute). In general, increased sympathetic activation
increases heart rate, while enhanced activation of parasympathetic, vagal tone decreases heart rate (Lahiri
et al.. 2008). HRV represents the degree of difference in the inter-beat intervals of successive heartbeats.
Given that both arms of the ANS contribute, changes in HRV are an indicator of the relative balance of
sympathetic and parasympathetic tone to the heart and their interaction (Rowan et al.. 2007). Low HRV is
associated with an increased risk of cardiac arrhythmia and an increased risk of mortality in patients with
CHF awaiting a heart or lung transplant (Fauchier et al.. 2004; Bigger etal.. 1992). Low HRV has also
been shown to be predictive of CAD (Kotecha et al.. 2012). Notably, increases in HRV have also been
associated with increases in mortality (Carll et al.. 2018). In general, the two most common ways to
measure HRV are time-domain measures of variability and frequency-domain analysis of the power
spectrum. With respect to time-domain measures, the standard deviation of normal-to-normal (NN)
intervals (i.e., the interval between consecutive normal beats) reflects overall HRV, and root-mean-square
of successive differences (rMSSD) in NN intervals reflects parasympathetic influence on the heart. In
terms of frequency domain, high-frequency (HF) domain is widely thought to reflect cardiac
parasympathetic activity while the low-frequency (LF) domain has been posited as an indicator of the
interaction of the sympathetic and parasympathetic nervous systems (Billman. 2013). although its linkage
with sympathetic tone is controversial and uncertain (Notarius et al.. 1999).
4.7.2.1 Epidemiologic Studies of Heart Rate and Heart Rate Variability
A small number of studies examining the relationship between Pb biomarkers and heart rate or
HRV were evaluated in the 2013 Pb ISA(U.S. EPA. 2013). However, the studies characterized in the
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2013 Pb ISA related to heart rate and HRV were all within the NAS cohort. Specifically, Park et al.
(2006) presented evidence of a relationship between patella Pb levels and decreased HRV among those
with three or more metabolic abnormalities (waist circumference >102 cm, hypertriglyceridemia
>150 mg/dL, HDL-C <40 mg/dL, BP >130/85 mmHg, fasting glucose >110 mg/dL). The results of this
study supported previous research presented in the 2006 Pb AQCD. Study-specific details, including
blood and bone Pb levels, study population characteristics, potential confounders, and select results from
these studies are presented in Table 4-11. These study details include standardized results as well as
results that could not be standardized based on the information provided in each paper.
In a more recent study (n = 203), Gump et al. (2017) evaluated the association between BLLs and
HRV among children (aged 9-11) as part of the Environmental Exposures and Child Health Outcomes
study. However, associations between BLLs (range: 0.19-3.25 (ig/dL) and HRV were null within this
group. Another recent analysis (n = 408) evaluated the effect of BLLs and HRV among children (age 12)
(Halabickv et al.. 2022).This study obtained blood Pb measurements at two time points (aged 3-5 and
12), while HRV was measured only at age 12. Children in this study were given a standardized stressful
stimulus known as the Public Speaking Stress task. In this task, children were asked to first plan a speech
to deliver (planning phase) and then present that speech to the research assistant (speaking phase) while
being continuously monitored for HRV. For the planning phase, there was a null association between a
HRV frequency measure (LF/HF) and BLLs at ages 3-5 (0.03 [-0.02, 0.09]) and at age 12 (-0.04 [-0.16,
0.07]). For the speaking phase, there was a positive association between HRV frequency and BLLs at
ages 3-5 (0.06 [0.01, 0.12]), but not with BLLs at age 12 (0.05 [-0.18, 0.08]). An increase in the LF/HF
ratio is associated with a shift to sympathetic dominance and an overall decrease in HRV, which is
suggestive of a dysregulated stress response.
4.7.2.2 Toxicological Studies of Heart Rate and Heart Rate Variability
The 2013 Pb ISA discussed a limited number of animal toxicological studies demonstrating that
exposure to Pb increased heart rate (SimSes et al.. 2011; Badavi et al.. 2008; Lai et al.. 2002). There were
no studies that examined changes in HRV in response to Pb exposure.
Since the publication of the 2013 Pb ISA, there have been additional toxicological studies
published with respect to exposure to Pb and HR. Fioresi et al. (2014) reported statistically significantly
higher heart rate (p < 0.05) in rats with a BLL of 13.6 (ig/dl, relative to control animals. However,
Wildemann et al. (2015) reported no change relative to control animals for heart rate in rats with a
1.7 (ig/dl or 8.6 |ig/dl BLL or following combined exposure with mercury or methylmercury. Other
studies were similarly mixed, with some reporting statistically significant increases in heart rate following
Pb exposure (Zhu et al.. 2019; Zhu et al.. 2018). while another study using two different exposure
scenarios did not (Shvachiv et al.. 2018). Thus, overall, there is mixed evidence from animal toxicological
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studies for an increase in heart rate following Pb exposure. Additional details for the toxicological studies
discussed above can be found in Table 4-12 of this ISA.
Since the 2013 Pb ISA, there have also been animal toxicological studies published with BLLs
<30 (ig/dl examining the relationship between Pb exposure and changes in HRV. Shvachiv et al. (2018)
reported a statistically significant increase in LF in rats continuously, but not intermittently exposed to Pb
relative to control animals. However, no changes in HF, or the LF/HF ratio were reported in either group
relative to controls. BLLs in this study were approximately 24 |ig/dl for the constant exposure group and
approximately 19 |ig/dl for the intermittent exposed group (Shvachiv et al.. 2018). Additionally, in a pair
of analyses by the same laboratory, there was a statistically significant increase (p < 0.05) in the LF/HF
ratio and a statistically significant decrease in LF and HF. BLLs in this study were <30 |ig/dl (Zhu et al..
2019; Zhu et al.. 2018). Thus, there is only limited evidence from animal toxicological studies for an
effect of Pb exposure on measures of HRV at BLLs <30 (ig/dl. Additional details for the toxicological
studies discussed above can be found in Table 4-12 of this ISA.
4.7.3 Integrated Summary of Cardiac Electrophysiology and Arrythmia
Exposure to Pb has been shown to affect contractility in animals and to be associated with cardiac
contractility in epidemiologic studies. The epidemiologic evidence supports an association between
altered ECG measures and biomarkers of Pb exposure. Specifically, a series of studies using the NAS
cohort presented in the 2013 Pb ISA indicated an association between bone Pb levels and a prolonged QT
interval (Eum et al„ 2011; Park et al.. 2009; Cheng et al.. 1998). There is evidence suggesting that a
lengthening of the QT interval increases risk of future abnormal heart rhythm or sudden cardiac arrest.
However, these NAS studies were small and evaluated mostly white, elderly men. A recent study
evaluated an earlier cohort of NHANES participants (NHANES III 1988-1994) (Jing et al., 2019). This
study included a much larger sample size and a more diverse group of subjects. In this cross-sectional
analysis, there was evidence of an increased QRS-T angle associated with BLLs. The effect was most
prominent in males compared with females. Despite the relatively consistent evidence observed within the
epidemiologic literature, the toxicological literature is sparce and more mixed. There was a single study in
the last review demonstrating increased incidence of arrhythmia, atrioventricular block, and a prolonged
ST segment interval (Reza et al„ 2008). Since the last review, an additional study reported no change in
the PR, QRS, or QT segments in rats (Wildemann et al„ 2015)
The epidemiologic evidence for an association between biomarkers of Pb exposure and either
heart rate or HRV are less compelling. Few studies evaluate this outcome. An earlier analysis of the NAS
cohort indicated an association between bone Pb measurements and a decrease in HRV among elderly
white men (Park et al.. 2006). This supported evidence from occupational studies presented in the 2013
Pb ISA (U.S. EPA, 2013). Results from recent studies in children are not consistent. An analysis among a
small group of children yielded no association between BLLs and HRV Gump et al. (2017), whereas a
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separate analysis indicated a slight decrease in HRV among 12-year-old children with their BLLs when
they were between the ages of 3-5 (Halabickv et al.. 2022). The toxicological evidence for exposure to Pb
and changes in heart rate was largely mixed. Some animal toxicological studies reported increases in heart
rate following Pb exposures (Zhu et al.. 2019; Zhu et al.. 2018; Fioresi et al.. 2014). whereas other animal
studies reported no change (Shvachiv et al.. 2018; Wildemann et al.. 2015). With respect to HRV, there
was a limited number of animal toxicological studies, but they reported changes in some measures of
HRV following Pb exposure (Zhu et al.. 2019; Shvachiv et al.. 2018; Zhu et al.. 2018). That said, there
were differences among these studies with respect to which measures of HRV changed, or the direction of
change for a given measure. Taken together, the relatively small body of evidence from epidemiologic
and toxicological studies examining Pb exposure and changes in cardiac electrophysiology and arrythmia
have reported mixed results.
4.8 Atherosclerosis and Peripheral Artery Disease
Atherosclerosis is the process of plaque buildup into lesions on the walls of the coronary arteries
that can lead to vessel narrowing, reduced blood flow to the heart, and IHD. The development of
atherosclerosis is dependent on the interplay between plasma lipoproteins, inflammation, endothelial
activation, and neutrophil attraction to the endothelium, extravasation, and lipid uptake. Risk factors for
atherosclerosis include high low-density lipoprotein (LDL) cholesterol/low HDL cholesterol, high BP,
diabetes, obesity, smoking, and increasing age. Measures of subclinical atherosclerosis provide the
opportunity to assess the pathogenesis of vascular disease at an earlier stage. PAD is an indicator of
atherosclerosis and is measured by the ankle brachial index, which is the ratio of BP between the posterior
tibia artery and the brachial artery. An ankle brachial index of less than 0.9 is typically indicative of the
presence of PAD. Prior toxicological studies have reported that Pb can increase atheromatous plaque
formation in pigeons, increase arterial pressure, decrease heart rate and blood flow, and alter cardiac
energy metabolism and conduction (Prentice and Kopp. 1985; Revis et al.. 1981).
4.8.1 Epidemiologic Studies of Atherosclerosis and Peripheral Artery
Disease
A limited number of studies have evaluated the effects of biomarkers of Pb exposure and
atherosclerosis. The 2013 Pb ISA (U.S. EPA, 2013) described an association between BLLs and both
intimal medial thickening (IMT) and atherosclerotic plaque presentation in an occupational study, among
those with high concentrations of blood Pb (-25 (ig/dL) (Poreba et al., 2011). Recent studies further
expand the knowledge base for the relationship between Pb biomarkers and atherosclerosis and PAD.
Study-specific details, including BLLs, study population characteristics, potential confounders, and select
results from these studies are highlighted in Table 4-13. These details include standardized results as well
as those that could not be standardized based on the information provided in each paper.
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A study published since the 2013 Pb ISA evaluated diabetic patients in China (Wan et al.. 2021).
This study evaluated the association between common carotid artery (CCA) plaques and BLLs. When
comparing the highest quartile of BLLs (>3.7 (ig/dL) with the lowest quartile of BLLs (<1.8 (.ig/dL). there
were increased odds (OR: 1.53 [95% CI: 1.29, 1.82]) of CCA plaque. The diameter of the CCA did not
appear related to BLLs (Figure 4-16).
Another recent analysis described the association between hemodynamic measures (peripheral
BP, central BP, and time-dependent hemodynamics), which assess arterial stiffness, and BLLs among a
Flemish population (Yu et al.. 2020). Blood Pb was collected at least once during the study period (1985
to 2005), and participants were followed for a median of 9.4 years. BLLs within this population were
relatively low (GM: 2.93 (ig/dL, IQR: 1.8-4.7). At the final follow-up, trained personnel assessed
measures of arterial stiffness. Overall, measures of peripheral BP or central BP were not associated with
BLLs. However, for every doubling of BLLs, several measures of time-dependent hemodynamics were
elevated, including augmentation ratio (1.74% [95% CI: 0.95, 2.53%]), augmentation index (3.03% [95%
CI: 1.56, 4.50]), forward pulse peak time (6.62% [95% CI: 2.21, 11.0%]), backward PP amplitude
(1.02 mmHg [95% CI: 0.02, 2.02 mmHg]), and reflection index (3.98% [95% CI: 1.71, 6.24%]).
However, the association with aortic pulse wave velocity (aPWV) was null (0.14 ms [95% CI: -0.08,
0.35 ms]), and age appeared to be a major component of increases in aPWV (Figure 4-18). The sum of
these results from this study indicates an association between relatively low BLLs and evidence of
atherosclerosis (Table 4-13).
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Panel A: pulse wave velocity was only standardized to a heart rate of 60 beats per minute; Panel B: the associations were fully
adjusted.
Source: Yu et ai. (2020).
Figure 4-18 Association between aortic pulse wave velocity with blood Pb
levels and age.
A large Korean study evaluated the association between BLLs and coronary artery stenosis
(CAS), which is the blockage or narrowing of the arteries that supply blood to the heart ( Cim et al..
2021). This study performed a coronary computerized tomography (CT) angiography and classified
participants with CAS if they had >25% stenosis. Overall, each 1 f ig/dL increase in BLL was associated
with increased odds of CAS (OR: 1.14 [95% CI: 1.02, 1.26]). Many studies of atherosclerosis focus on
calcification or blockage of the coronary artery, but a recent NHANES analysis focused on abdominal
aortic calcification (AAC) (Din et al.. 2021). AAC is a marker of subclinical atherosclerosis and a
predictor of future CYD events. Lateral lumbar spine images, using the Kauppila score system were used
to score the AAC severity on a scale from 0 to 24, and a total AAC score >6 was considered to be
substantial calcification of the abdominal aorta. Overall, each one-unit increase in BLLs corresponded to
a 0.15-unit increase (95% CI: 0.02, 0.27) in total AAC score and an 11% increase in severe AAC (OR:
1.11 [95% CI: 1.00, 1.22]). There were no differences in association when stratified by race, sex, age,
BMI, hypertension, or diabetic status (Figure 4-19).
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AAC Score
P (95% CI)
P for trend
P for interaction
Race
Mexican American
-0.02 (-0.28.0.24)
0.9
0.49
Other Hispanic
0.37 (-0.05. 0.78)
0.084
Non-Hispanic White
0.34 (0 09. 0.59)
0.0075
Non-Hispanic Black
-0.07 (-0.22. 0.09)
0.4
¦—•—>
Other Races
0.01 (-0.33. 0.34)
098
Gender
Male
0.16(0.02. 0.30)
0.025
•—•—i
0.22
Female
0.27 (0.17.0.37)
0.045
-¦-
Age (years)
Age < 60
0.11 (0.00.0.22)
0.031
0.34
Age 2 60
0.23 (0.16. 0.29)
0.042
BMI
Normal weight
0.05 (-0.14, 0.38)
0.36
0.61
Overweight
0.09 (0.01.0.22)
0.0019
>-•—
Obese
0.32(0.12, 0.52)
0.0023
Hypertension
Yes
0 17(0 09, 0 4?)
n n?i
i-m ¦
0 64
No
0.13(0.00. 0.25)
0 044
Diabetes
Yes
0.05 (0.01.0.12)
0.02
¦ <
0.33
NO
0.16(0.04. 0.28)
0.01
1 1
1
AAC = abdominal aortic calcification; BLL = blood lead level; BMI = body mass index; CI = confidence interval; P = p-value.
Source: Qin et al. (2021).
Figure 4-19 Stratified associations between abdominal aortic calcification
score and blood Pb levels.
The 2013 Pb ISA described epidemiologic studies assessing the relationship between biomarkers
of Pb exposure and prevalent PAD. An NHANES (1999-2002) analysis observed an increasing trend in
the odds of PAD with increasing concurrent BLLs (Muntner et al.. 2005). Another NHANES (1999-
2000) analysis also indicated a trend of increasing odds of PAD with increasing quartiles of concurrent
BLLs, among adults >40 years (Navas-Acien et al.. 2004). However, these results were not statistically
significant for any quartile of Pb exposure. To date, no recent studies evaluating biomarkers of Pb
exposure and PAD have been conducted for inclusion in the current review.
4.8.2 Toxicological Studies of Atherosclerosis
In the 2013 Pb ISA, a study in rats demonstrated that Pb exposure increased the aortic media
thickness, media-lumen ratio, and medial collagen content (Zhang et al.. 2009). Since the publication of
that document, Xu et al. (20151 reported a statistically significant increase in proliferating cell nuclear
antigen in cardiac tissue (p < 0.05) in rats exposed to Pb up to 12 or 40 days. This result potentially
indicates increased cellular division and/or DNA repair following exposure to Pb, which is relevant given
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that increased cellular proliferation plays a role in atherosclerotic plaque growth. Moreover, in the 40-day
exposure group, these authors reported a statistically significant increase in the diameter of the cells of the
aorta, as well as changes in the shape (i.e., loss of curvature) of the aortic internal elastic lumen relative to
control animals. The blood Pb concentrations in this study were 19.3 (ig/dl on day 12 and 24.6 |ig/dl on
day 40 (Xu et al.. 2015). Additional details for the animal toxicological studies discussed in this section
can be found in Table 4-14 of this ISA.
4.8.3 Integrated Summary of Atherosclerosis
At the time of the 2013 Pb ISA, there were few studies examining the relationship between
biomarkers of Pb exposure and measures of atherosclerosis and PAD. Overall, these studies were mixed,
with some indicating an association between BLLs and IMT or an increase in the odds of PAD
prevalence, while others did not indicate a relationship between BLLs and prevalent PAD. Recent
epidemiologic evidence indicates a consistent positive association between markers of atherosclerosis and
Pb exposure. Atherosclerotic evidence is measured differently between the included studies, but further
supports the notion of an association between BLLs and plaque formation. While there is strong evidence
that markers of atherosclerosis increase with age, BLLs also appear to play a substantial role. The
toxicological evidence from the previous and current ISA is limited but supports epidemiologic studies
demonstrating a positive association between Pb exposure and markers of atherosclerosis. More
specifically, there is animal toxicological evidence of morphological changes in the aorta consistent with
the potential for atherosclerosis. Taken together, there is evidence from both epidemiologic and
toxicological studies to support an association between biomarkers of Pb exposure and makers of
atherosclerosis development.
4.9 Cerebrovascular Disease
Cerebrovascular disease describes a group of conditions involving the cerebral blood vessels that
result in transient or permanent disruption of blood flow to the brain. These conditions include stroke,
transient ischemic attack, and subarachnoid hemorrhage. Both hypertension and atherosclerosis are risk
factors for cerebrovascular disease and the mechanisms for these outcomes also apply to cerebrovascular
disease. Very few studies have examined the effects of Pb exposure on cerebrovascular disease.
4.9.1 Epidemiologic Studies of Cerebrovascular Disease
The 2013 Pb ISA (U.S. EPA, 2013) described a limited number of epidemiologic studies that
examined associations between Pb exposure and cerebrovascular disease. Two previous prospective
epidemiologic studies evaluated mortality from stroke(Khalil et al., 2009; Menke et al„ 2006). In an
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NHANES analysis, Menke et al. (2006) indicated that increases in BLLs were associated with an increase
in stroke mortality, although the association was imprecise. In contrast, Khalil et al. (2009) reported a null
association between BLLs and stroke mortality. In a cross-sectional study in Taiwan, Lee et al. (2009)
reported an association between increased intracranial and extracranial stenosis (>50%) and urine Pb
concentrations but not blood Pb concentrations.
In a recent small case-control analysis (n = 88), Mousavi-Mirzaei et al. (2020) evaluated acute
ischemic stroke in relation to BLLs among patients in Iran. Cases (n = 44) of acute ischemic stroke were
matched to controls (n = 44) based on age, sex, occupation, opium addiction, and sampling time.
Participants in this study had relatively high BLLs (median: 6.38 (ig/dL, IQR: 1.75-34.87). There was an
association between increased BLLs and increased risk of acute ischemic stroke (OR: 1.04 [95% CI: 1.02,
1.07] for a 1 (ig/dL increase in blood Pb). Study-specific details, including BLLs, study population
characteristics, potential confounders, and select results are highlighted in Table 4-15. Study details in
Table 4-15 include standardized results as well as results that could not be standardized based on the
information provided in each paper There were no animal toxicological studies at BLLs <30 (ig/dL that
examined the relationship between Pb exposure and cerebrovascular disease.
4.9.2 Summary of Cerebrovascular Disease
Few studies have examined the relationship between biomarkers of Pb exposure and
cerebrovascular disease. A small amount of evidence was presented in the 2013 Pb ISA suggesting an
association between Pb exposure and stroke mortality or stenosis in the intracarotid system. Since the
publication of these prior documents, however, very little additional epidemiologic information can be
added to the current evidence base. Moreover, there were no relevant animal toxicological studies at
BLLs <30 (ig/dL. Thus, the evidence to suggest an association between Pb exposure and cerebrovascular
disease is limited.
4.10 Cardiovascular Mortality
4.10.1 Epidemiologic Studies of Cardiovascular Mortality
Studies that examine the association between biomarkers of Pb exposure and cause-specific
mortality outcomes, such as cardiovascular mortality, provide additional evidence for Pb-related
cardiovascular effects, specifically whether there is evidence of an overall continuum of effects. Several
epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA, 2013) strengthened the evidence
presented in the 2006 Pb AQCD (U.S. EPA, 2006) indicating an association between Pb biomarkers of
exposure and cardiovascular mortality. The strongest evidence came from multiple prospective cohort
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studies that observed consistent positive associations with CVD mortality across different populations,
while also using different model specifications and approaches to control for a wide range of potential
confounders. The majority of cohort studies evaluated in the 2013 Pb ISA utilized blood Pb data from
NHANES II and III, which was then linked prospectively to mortality data, with between 8-16 years of
follow-up (Mcnkc et al.. 2006; Schober et al.. 2006; Lustberg and Silbergeld. 2002). Additional
prospective cohort studies, specifically among older adults, reported that CVD mortality was associated
with Pb measured in blood (Khalil et al.. 2009).
Notably, adult BLLs may be representative of contributions from both recent Pb exposures and
mobilization of legacy Pb from bone, therefore it remains unclear as to what extent either recent, past, or
cumulative Pb exposures contribute to the observed associations with cardiovascular mortality. Because
of the rapid decline in ambient air Pb concentrations and population BLLs that corresponded with the
phase out of leaded gasoline, participants of NHANES II (1976-1980) and NHANES III (1988-1994)
likely had higher past Pb exposures compared with exposure at the time of blood collection—further
complicating the determination of BLLs that might contribute to the observed associations. Recent studies
continue to provide evidence of consistent positive associations between exposure to Pb and CVD
mortality (Figure 4-20). Study-specific details, including biomarker Pb levels, study population
characteristics, confounders, and select results from these studies, are highlighted in Table 4-16.
Reference Study
Cardiovascular Mortality:
Population Pb distribution
Menke et al.,2006
fLanphearetal., 2018
NHANES III Adults £20
tDuan et al., 2020*
NHANES Adults £ 20
Median (IQR)
1.49(0.93, 2.31)
Cause-specific Cardiovascular Mortality:
Pb measurement -Years of
NHANES III Adults £ 20 Mean: 2.59
Geometric Mean: 2.71
Geometric SE: 1.32
Median
fVan Bemmel et al., 2011 NHANES III Adults £40 <5ug/dL2.6
a 5 ug/dL 7.5
1988-1994
1988-1994
12
19
NA
CVD
CVD ALAD GG
CVD ALAD CG/GG
Menke et al., 2006
Menke et al., 2006
TLanphear et al., 2018 NHANES III Adults £20
NHANES III Adults £ 20 Mean: 2.60
NHANES III Adults £20 Mean: 2.61
Geometric Mean: 2.71
Geometric SE: 1.33
1988-1994
1988-1994
12
19
Stroke
IHD
0.90 1.00 1.10 1.20
Effect Estimate (95% CI) per 1 ug/dL increase in blood Pb
ALAD = 6-aminolevulinic acid dehydratase; ALAD GG and ALAD CG/GG = variants of 6-aminolevulinic acid dehydratase;
CI = confidence interval; CVD = cardiovascular disease; IHD = Ischemic heart disease; IQR = Interquartile range Ml = myocardial
infarction; NHANES = National Health and Nutrition Examination Survey; Pb = lead; T# = fertile #.
Note: fRed text: Studies published since the 2013 Pb ISA, Black text: Studies included in the 2013 Pb ISA.
Effect estimates are standardized to a 1 |jg/dL increase in blood Pb. If the Pb biomarker is log-transformed, effect estimates are
standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed
to be linear within the evaluated interval.
*Study estimated relative risk.
Figure 4-20 Associations between blood Pb level and cardiovascular
mortality.
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In an analysis of the NHANES III cohort, Lanphear et al. (2018) reported that a 1 (ig/dL increase
in BLLs was associated with hazard ratios (HRs) of 1.10 (95% CI: 1.05, 1.15]) for CVD mortality and
1.14 [95% CI: 1.08, 1.20]) for IHD mortality. Lanphear et al. (2018) extended the average follow-up time
of the Menke et al. (2006) analysis of the same NHANES III cohort by over 7 years (from ~12 to
-19 years), resulting in a substantial increase in observed cardiovascular deaths (766 versus 1,801).
Several other recent studies that analyzed NHANES cycles reported associations of similar magnitude for
CVD mortality (Duan et al.. 2020; Ruiz-Hernandez et al.. 2017; Aoki et al.. 2016; van Bemmel et al..
2011). Specifically, van Bemmel et al. (2011). Ruiz-Hernandez et al. (2017). and Cook et al. (2022)
assessed cohorts using NHANES III data with similar results. Aoki et al. (2016) evaluated the relationship
between CVD mortality and BLLs with additional control for either hemoglobin or hematocrit values
using NHANES (1999-2010) data with mortality follow-up through 2011. A 10-fold increase in BLLs
was associated with an RR of 1.26 (95% CI: 0.91, 1.78). However, when BLLs were hemoglobin-
corrected (see details below), there was a greater increase in magnitude and precision in predicting CVD
mortality, compared with the association with whole blood Pb alone (RR: 1.46 [95% CI: 1.06, 2.01]).
Similar results were obtained when evaluating hematocrit-corrected whole BLLs and CVD mortality (RR:
1.44 [95% CI: 1.05, 1.98]).
Duan et al. (2020) evaluated the relationship between CVD mortality and BLLs using NHANES
(1999-2014) with mortality follow-up through 2015. Here, a 1 (ig/dL increase in BLLs was associated
with an RR of 1.39 (95% CI: 1.28, 1.51). Both Aoki et al. (2016) and Obeng-Gvasi et al. (2021) also
relied on more recent NHANES blood Pb data, 1999-2010 and 1999-2008, respectively. Because of the
phaseout of leaded paint and gas, these more recent NHANES studies will capture populations potentially
less affected by the earlier period of elevated Pb exposure. . It is expected, however, that these
populations would still have had a substantial period of elevated BLLs in early life due to the gradual
decline in BLLs overtime.
A number of studies have additionally evaluated either hemoglobin- or hematocrit-corrected
BLLs and mortality. Aoki et al. (2016) used six 2-year NHANES cycles (1999-2010) linked with
mortality data through the end of 2011 (median follow-up time: 6.2 years) to evaluate the association
between whole BLLs, hematocrit-corrected and hemoglobin-corrected BLLs, and CVD mortality among
subjects >40 years of age at baseline. Hematocrit- or hemoglobin-corrected whole blood Pb was
calculated by dividing whole blood Pb by either hematocrit or hemoglobin, respectively. To make the
results more comparable with whole blood Pb, the values were multiplied by the weighted arithmetic
mean of either hematocrit or hemoglobin (Aoki et al.. 2017). In models assessing whole BLLs (not
corrected for either hematocrit or hemoglobin), every 10-fold increase in whole BLLs was associated with
an RR of 1.26 (95% CI: 0.91, 1.78) for cardiovascular mortality. Results were similar when controlling
for hematocrit (RR: 1.35 [95% CI: 0.98, 1.86]) or hemoglobin (RR: 1.35 [95% CI: 0.98, 1.87]) as a
covariate in the model. However, the association was stronger in terms of both magnitude and precision
when evaluating hematocrit-corrected (RR: 1.44 [95% CI: 1.05, 1.98]) or hemoglobin-corrected (RR:
1.46 [95% CI: 1.06, 2.01]) whole BLLs. Another study (Lin et al.. 2011). also examined BLLs corrected
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for hemoglobin. Lin et al. (2011) examined Taiwanese adults with end-stage renal disease with relatively
high (mean: 11.5 (ig/dL) BLLs. To correct for hemoglobin, the authors used the following equations, for
males: BLL x 14/hemoglobin concentration; for females: BLL x 12/hemoglobin concentration. The
hemoglobin-corrected blood Pb results were similar in magnitude (HR: 7.35 [95% CI: 1.64, 33.33]) than
the noncorrected blood Pb values (HR: 9.71 [95% CI: 2.11, 23.26]) when comparing the highest tertile
(>12.64 (ig/dL) with the lowest tertile (<8.51 (.ig/dL).
A recent re-analysis of NAS data (Weisskopf et al.. 2015). expanded on a similar analysis
(Weisskopf et al.. 2009) which was presented in the 2013 Pb ISA. In the re-analysis, special
considerations for selection bias were taken. Specifically, the authors created four different models, which
controlled for different covariates, additional markers for SES, and restricted by age (Table 4-16). In this
analysis, the authors restricted the sample (Model 3 and Model 4) to participants that were <45 years at
the start of the NAS study, since cardiovascular disease-related deaths would be relatively rare in the
younger population. This study indicated a positive association with CVD (HR: 2.23 [1.02, 4.84]) and
IHD (HR: 4.60 [1.26, 16.8]) when comparing the highest tertile (>31 (ig/g) of patella Pb to the lowest
tertile (<20 |ig/g). in the model restricting the age of participants to participants <45 years at the start of
the NAS study. No associations were observed without the age restriction or with blood or tibia Pb.
In a recent study, Hollingsworth and Rudik (2021) implemented a quasi-experimental design to
examine the effect of the phase out of leaded gasoline in automotive racing on mortality rates in older
adults. Comparing time periods prior to and after the phaseout of leaded gasoline in professional racing
series (i.e., the National Association for Stock Car Auto Racing [NASCAR] and the Automobile Racing
Club of America [ARCA]), the authors used a difference-in-differences technique to estimate county-
level changes in air Pb concentrations, elevated BLL prevalence among children, and mortality rates in
race counties and counties bordering race counties relative to control counties. A detailed discussion of
results for air Pb concentrations and BLLs is presented in Appendix 2. Section 2.4.1. In short, there were
substantial declines in both air Pb concentrations and the prevalence of children with elevated BLLs
associated with the phaseout of leaded gasoline. The authors also reported significant declines in
cardiovascular mortality rates over this same period. Specifically, in the period following de-leading of
gasoline, there was an estimated decline in annual age-standardized cardiovascular mortality rates of 37
deaths per 100,000 in race counties and 12 deaths per 100,000 in border counties. Additionally, there was
a similar decline for IHD-related deaths, with 53 deaths per 100,000 in race counties and 20 deaths per
100,000 in border counties. Similar to the exposure results, the mortality estimates appear to demonstrate
a distance gradient. The difference-in-difference approach controls for spatially varying confounders by
estimating the difference in mortality rates in adjoining years in the same county and controls for
temporally varying confounders by taking the difference of those differences between locations. The
authors additionally adjust for potential confounders that may vary spatially and temporally
(e.g., unemployment rate and quantity of Toxic Release Inventory [TRI] lead emissions). Hollingsworth
and Rudik (2021) did not adjust for potential co-pollutant exposures but provides evidence that there is no
differential effect of leaded and unleaded races on other co-pollutant concentrations (i.e., CO, VOCs,
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PMio, PM2.5, NO2, and O3) in the weeks leading up to and following the race. However, because the
mortality rates are an annual measure, there is remaining uncertainty regarding potential differential
trends in the long-term average of other pollutants that could be correlated with the phaseout of leaded
gasoline in NASCAR and ARCA.
Several analyses evaluated metal chelation therapy as a treatment for those with atherosclerotic
plaques and evaluated subsequent mortality outcomes in the Trial to Assess Chelation Therapy (TACT)
study. The TACT study was a randomized control trial (RCT) with a 2 x 2 factorial design evaluating
chelation therapy with ethylenediaminetetraacetic acid (EDTA) plus the use of high dose oral vitamins.
The factorial group results indicated that a combination of EDTA and high-dose vitamins was associated
with a reduction in clinically important cardiovascular events, especially for cardiovascular deaths, MI, or
stroke (Lamas et al.. 2014). In the same trial, the findings indicated that diabetic patients >50 years, had a
reduction (6% versus 9% HR: 0.63 [95% CI: 0.35, 1.13]) in cardiovascular death following EDTA
chelation therapy (Escolar et al.. 2014). Although these studies suggest a clear association between
chelation therapy and a reduction in cardiac deaths, these studies did not measure BLL pre and post
chelation making the potential role of Pb unclear, as compared to other divalent ions that are chelated by
EDTA.
Additionally, a recent meta-analysis (Chowdhurv et al.. 2018) evaluated several metal
biomarkers, including BLLs, and evaluated the overall association between BLLs and CHD, specifically
including studies of CHD mortality. It included studies with cohort, case-control, or nested-case-control
study designs. As described in Section 4.4, this analysis provided evidence of an increased risk of CHD
mortality associated with increasing BLLs (Figure 4-20).
4.10.1.1 Dose-Response Relationship
An examination of the dose-response relationship between biomarkers of Pb exposure and
cardiovascular mortality helps to further evaluate the continuum of effect between biomarkers of Pb
exposure and cardiovascular outcomes. Because of differences in exposure historically, it is expected that
adult BLLs would be influenced by historical Pb exposures. Therefore, studies examining a single blood
Pb measurement in adulthood may not fully capture the true effect of biomarkers of Pb exposure and
cardiovascular mortality.
Several recent studies, however, have summarized mortality outcomes over a range of blood Pb
values. Lanphear et al. (2018) extended the follow-up period of the Menke et al. (2006) study and
evaluated the dose-response relationship among the same population. Using a five-knot restricted cubic
spline analysis, this study generally indicated a supralinear dose-response relationship between BLLs and
CVD and IHD mortality. The authors also stated that this dose-response relationship was steeper (HRs
were larger in magnitude) at lower blood Pb concentrations. Overall, this study reported increased risk of
CVD or IHD mortality among those with BLLs <5 (ig/dL (Figure 4-21).
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Concentration of lead in blood (ng/dL)
CI = confidence interval.
Note: Restricted cubic spline (5 knots) (red lines) and adjusted HRs (black lines) with 95% CIs (hatched lines) for (B) cardiovascular
disease mortality and (C) IHD mortality.
Source: Adapted from Lanphear et al. (2018).
Figure 4-21 Dose-response relationship between blood Pb levels and
cardiovascular and ischemic heart disease mortality.
These results are similar to other assessments of the dose-response relationship described in
previous assessments of Pb, including the 2006 Pb AQCD (U.S. EPA, 2006) along with the 2013 Pb ISA
(U.S. EPA, 2013). In the original evaluation of the NHANES III data and mortality, Menke et al. (2006)
noted a similar linear shape of the dose-response curve. Specifically, the dose-response relationship was
steeper at lower blood Pb concentrations.
A similar NHANES III (1988-1994) analysis evaluated total CVD, heart disease (CVD diagnosis
codes excluding stroke), and MI deaths through 2010 (Cook et al., 2022). This study indicated a greater
risk of CVD mortality among those with the highest BLLs (>6.23 (tg/dl for men and >3.74 (tg/dl for
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women) (Figure 4-22). Similar patterns were reported for heart disease and MI mortality. The results of
this analysis provide no evidence of a threshold below which an association between blood Pb and
mortality does not exist, at least within the blood Pb ranges within this study (10th percentile: 1.0 (ig/dl,
90th percentile: 6.7 (ig/dl). A sensitivity analysis evaluated BLLs continuously. In the model, a 1-unit
increase in log-transformed BLLs was associated with an 8% (HR: 1.08 [95% CI: 1.00, 1.16]) increase in
total CVD mortality risk and a 9% (HR: 1.09 [95% 1.02, 1.16]) increase in heart disease risk. However,
there was no reported increased risk of acute MI mortality associated with a 1-unit increase in log-
transformed BLLs (HR: 0.95 [95% CI: 0.84, 1.08]).
CIF = cumulative incidence function; CVD = cardiovascular disease; NHANES = National Health and Nutrition Examination Survey.
Source: Cook et al. (2022).
Figure 4-22 Cumulative incidence function of cardiovascular mortality by
blood Pb level, National Health and Nutrition Examination Survey
III (1988-1994).
4.10.1.2 At-Risk Populations
Several recent analyses of biomarkers of Pb exposure and cardiovascular mortality have
evaluated EMM or stratification of the relationship by specific parameters such as sex, genetic factors,
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stress, and behavior factors like smoking, whereas other analyses have primarily focused on populations
or lifestages that may be particularly vulnerable to premature mortality associated with biomarkers of Pb
exposure. The differences observed in at-risk populations are described below.
4.10.1.2.1 Effect measure modification
Ruiz-Hernandez et al. (2017) used NHANES III (1988-1994) and three 2-year NHANES cycles
(1999-2004) linked with mortality data—through the end of 1996 for the NHANES III cohort and
through the end of 2006 for the NHANES 1999-2004 cohort data—to assess the relationship between
BLLs and CVD and CHD mortality. This study showed that in fully adjusted models, there were
increases in both CVD mortality (relative risk [RR]: 1.19 [95% CI: 1.07, 1.31]) and CHD mortality (RR:
1.24 [95% CI: 1.10, 1.41]) for each doubling of BLLs. The RRs for both CVD and CHD mortality were
stronger among women compared with men and among never-smokers compared with ever-smokers.
Despite a higher mean BLL in the NHANES 111(1988-1994) cohort, the RR was 1.17 (95% CI: 1.06,
1.29) compared with 1.43 (95% CI: 1.16, 1.78) in 1999 to 2004.
van Bemmel et al. (2011) investigated a smaller subset of NHANES III (1988-1994) subjects.
This study specifically evaluated EMM of the relationship between BLLs and cardiovascular mortality by
polymorphisms in ALAD. A critical mechanism of Pb toxicity is its ability to interact and inhibit key
enzymes, such as ALAD, in the heme biosynthetic pathway. This analysis identified a null association
between elevated BLLs (>5 (ig/dL) and cardiovascular mortality. When further stratified by ALAD
variant, this study continued to observe null associations of elevated BLLs (>5 (ig/dL) among both
ALADGG variants for cardiovascular mortality (HR: 1.01 [95% CI: 0.92, 1.10]) and among ALADCG/CC
variants (HR: 1.13 [95% CI: 0.93, 1.36]).
In a more recent analysis, (Obeng-Gvasi et al.. 2021) evaluated whether the association between
BLLs and CVD mortality was modified by AL, a measure of cumulative stress. This study used
NHANES (1999-2008) data linked to mortality data through 2014. First, the study indicated that higher
BLLs were associated with a higher AL index. There was also an increased risk of CVD mortality among
those with BLLs >1.55 (ig/dL (median) (HR: 2.35 [95% CI: 1.77, 2.93]), when compared with those
below the median BLL. This study also indicated that the interaction between BLLs and AL was
significant (p = 0.014) but did not present stratified results.
Evidence of EMM is in direct contrast to stratified analyses presented in the previous Pb ISA.
Menke et al. (2006) demonstrated that there were no interactions between BLLs and other adjusted
variables, when comparing the 80th percentile (4.92 (ig/dL) with the 20th percentile (1.46 (ig/dL) of
BLLs. Specifically, the association between BLLs and cardiovascular mortality was positive but not
different when stratified by age, race, sex, urban/rural residence, smoking, BMI, and comorbid conditions
(hypertension, diabetes, low kidney function). However, Lanphear et al. (2018) indicated EMM by age, in
the same NHANES population, but with longer follow-up. Specifically, individuals >50 years old had
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greater HRs for cardiovascular disease mortality (HR: 2.93 [95% CI: 1.60, 5.36] vs. HR: 2.08, [95% CI:
1.35, 3.19]) and for IHD mortality (HR: 4.68 [95% CI: 2.42, 9.05] vs. HR: 2.46, [95% CI: 1.51, 4.01]).
Additionally, the HR for cardiovascular disease mortality was higher in smokers (HR: 2.19 [95% CI:
1.47, 3.26]), compared to non-smokers (HR: 1.32 [95% CI: 0.86, 2.05]).
4.10.1.2.2 Specific Populations
Several recent analyses have focused on the analysis of biomarkers of Pb exposure and
cardiovascular mortality among certain populations with comorbid conditions or at specific lifestages. Lin
et al. (2011) evaluated the relationship between BLLs and mortality among patients on maintenance
hemodialysis in a relatively short (-18 months of follow-up) prospective cohort in Taiwan. Study subjects
had a relatively high average BLL (mean: 11.5 (.ig/dL). which is higher than the general Taiwanese
population (mean: 7.7 (ig/dL). It is suspected that hemodialysis patients may experience higher BLLs
because their kidneys may no longer be able to excrete Pb from the body due to a total loss of renal
function (see Appendix 5: Renal Effects). There was an increased HR among those in the third tertile of
BLLs (>12.64 (ig/dL) for cardiovascular mortality (HR: 9.71 [2.11, 23.26]), compared with the first tertile
of BLLs (<8.51 (ig/dL). Additionally, when considering hemoglobin-corrected blood Pb values, the
association between the highest and lowest tertile was smaller in magnitude (HR: 4.98 [95% CI: 1.86,
13.33]) compared with whole blood Pb measurements, but was still imprecise (both measurements had
large confidence intervals).
4.10.2 Summary of Cardiovascular Mortality
The CVD mortality results in this review supported and expanded on findings from both the 2006
Pb AQCD, which included NHANES mortality studies (Schober et al.. 2006; Lustbcrg and Silbcrgcld.
2002). and the 2013 Pb ISA, which included an NHANES mortality study (Menke et al.. 2006) and non-
NHANES cohort analyses (Khalil. 2010; Khalil et al.. 2009; Weisskopf et al.. 2009). Several of the most
recent NHANES analyses (Duan et al.. 2020; Lanphear et al.. 2018; Ruiz-Hernandez et al.. 2017; Aoki et
al.. 2016) further strengthen the evidence provided within the 2013 Pb ISA, by including a wide range of
potential confounders and further consideration of a dose-response relationship. Furthermore, the most
recent NHANES analyses provide evidence of an association between BLLs and mortality at lower mean
population blood Pb concentrations (mean or median blood Pb range between 1.49 and 3.2 (ig/dL) (Duan
et al.. 2020; Lanphear et al.. 2018; Ruiz-Hernandez et al.. 2017; Aoki et al.. 2016). Despite the differences
observed within the studies, associations between increased concentrations of Pb biomarkers and
mortality were generally observed (Figure 4-23, Table 4-16).
There still remains uncertainty regarding the relative contributions of recent, past, and cumulative
Pb exposure for the relationship between BLLs and cardiovascular mortality. The more recent NHANES
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analyses evaluate cycles as recent as 2015 and continue to observe strong associations between
increasingly lower levels of blood Pb and cardiovascular mortality; however, these analyses still contain
populations greatly influenced by high historic Pb exposure. Additionally, further confounder control,
such as the inclusion of Cd concentrations in blood or urine, can also reduce the uncertainty noted in the
2006 Pb AQCD. van Bemmel et al. (2011) reported null associations between BLLs and cardiovascular
mortality. However, despite using data from NHANES III, the authors were not able to sufficiently
account for all confounders, and were limited to a smaller sample size, given their study hypothesis. The
cohort of Taiwanese hemodialysis patients provided evidence that there may be subsets of the population
at an increased risk of Pb-related cardiovascular mortality, compared with the general population (Lin et
al.. 2011). Taken together, despite differences in the design, methods, and considerations across studies,
associations between elevated levels of Pb biomarkers and increased mortality risk were generally
observed.
4.11 Biological Plausibility
Sections 4.1 to 4.10 of this appendix describe the cardiovascular health effects associated with
exposure to Pb from epidemiologic and animal toxicological studies. Informed largely by the animal
toxicological evidence presented in these sections, as well as in previous ISAs and AQCDs, this section
describes the biological pathways that potentially underlie the cardiovascular associations observed in
epidemiologic studies. Figure 4-23 graphically depicts these proposed pathways as a continuum of
pathophysiological responses—connected by arrows—that may ultimately lead to the apical
cardiovascular events associated with exposures to Pb at concentrations observed in epidemiologic studies
(e.g., IHD, MI). Note that the role of biological plausibility in contributing to the weight-of-evidence
causality determinations reached in the current Pb ISA are discussed in Section 4.12.
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Pb Exposure
Pro-
Atherosclerotic
Environment
(e.g., increased
inflammation,
clotting factors,
cholesterol)
Thrombosis
¦
¦
Exacerbation
of Ischemic
Heart Disease/
Potential
Myocardial
infarction or
Stroke
Cardiovascular
Related
Mortality
Modulation of
the Automatic
Nervous
System (e.g.,
HRV, HR)
HR = heart rate; HRV = heart rate variability; Pb = lead.
Note: The boxes above represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and
the arrows indicate a proposed relationship between those effects. Shading around multiple boxes is used to denote a grouping of
these effects. Arrows may connect individual boxes, groupings of boxes, and individual boxes within groupings of boxes.
Progression of effects is generally depicted from left to right and color-coded (gray, exposure; green, initial effect; blue, intermediate
effect; orange, effect at the population level or a key clinical effect). Here, population-level effects generally reflect results of
epidemiologic studies. The structure of the biological plausibility sections and the role of biological plausibility in contributing to the
weight-of-evidence analysis used in this Pb ISA are discussed in Section 4.12.
Figure 4-23 Potential biological pathways for cardiovascular effects following
exposure to Pb.
Considering the available health evidence, Figure 4-26 shows plausible pathways connecting Pb
exposure to the apical events reported in epidemiologic studies. The first potential pathway begins with
oxidative stress leading to impaired vascular function, systemic inflammation, a pro-atherosclerotic
environment, and increases in BP. The second potential pathway involves Pb perturbation of the RAAS
leading to increases in BP and impaired vascular function. The third potential pathway involves
modulation of the ANS leading to increases in BP and exacerbation of conduction abnormalities and
arrythmia. Once these pathways are initiated, there is evidence from in vitro and in vivo toxicological
studies that exposure to Pb may result in a series of pathophysiological responses that could lead to
cardiovascular events such as IHD, MI, and stroke, and thus, possible cardiovascular mortality.
As noted above, one potential pathway for Pb exposure to result in the associations reported in
epidemiologic studies is through the induction of oxidative stress and inflammation. Exposure to Pb can
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stimulate the production of ROSs in the blood, heart, and/or vasculature (SimSes et al.. 2015; Dewaniee et
al.. 2013; Farmand et al.. 2005; Ni et al.. 2004; Attri et al.. 2003; Courtois et al.. 2003; Vaziri et al.. 1999;
Gonick et al.. 1997) For example, Pb exposure in rats resulted in increased levels of superoxides and
hydrogen peroxide in human coronary endothelial cells. Similarly, Vaziri et al. (1999) demonstrated
increased plasma and cardiac levels of the oxidative stress marker 3-nitrotyrosine following Pb exposure
in rats.
Pb exposure resulting in the production of ROSs is important because several studies have
demonstrated a role for Pb-induced oxidative stress in impaired vascular function (SimSes et al.. 2015;
Dursun et al.. 2005; Attri et al.. 2003; Gonick etal.. 1997; Vaziri et al.. 1997; Khalil-Manesh et al.. 1994;
Khalil-Manesh et al.. 1993). Impaired vascular function is often the result of impaired functioning of the
endothelium, which maintains the normal balance of mediators that promote vasorelaxation (e.g., nitric
oxide) and vasoconstriction (e.g., endothelin-1). Animal toxicological studies demonstrate that exposure
to Pb results in altered vascular function (Nunes et al.. 2015; Gaspar and Cordellini. 2014; Silveira et al..
2010; Zhang et al.. 2007; Skoczvnska and Stoiek. 2005). including impairment that would be consistent
with greater vasoconstriction (SimSes et al.. 2015; Gaspar and Cordellini. 2014) or decreased vasodilation
(Silveira et al.. 2010; Zhang et al.. 2007). Toxicological studies have also demonstrated that Pb-induced
oxidative stress results in impaired vascular function through the inactivation or downregulation of
vasodilators such as nitric oxide and/or soluble guanylate cyclase, thereby increasing the potential for
vasoconstriction (Goncalves-Rizzi et al.. 2016; Dursun et al.. 2005; Attri et al.. 2003; Gonick et al.. 1997;
Vaziri et al.. 1997; Khalil-Manesh et al.. 1994; Khalil-Manesh et al.. 1993). For example, Pb-induced
ROSs can inactivate or sequester the vasodilator nitric oxide (Malvezzi et al.. 2001; Vaziri et al.. 1999).
and inhibition of nicotinamide adenine dinucleotide phosphate oxidase was able to block Pb-enhanced
contraction of cultured rat aorta cells in response to the vasoconstrictor 5 hydroxytryptamine (Zhang et
al.. 2005).
Continuing along this potential pathway, impaired vascular function also promotes plaque
formation potentially leading to atherosclerosis. In addition to its hallmark feature of impaired
vasodilation, impaired vascular function is further characterized by decreased vascular integrity, increased
expression of adhesion molecules, and cytokine upregulation (Lind et al.. 2021). In total, this increases
the potential for atherosclerotic disease and formation of thrombi (i.e., blood clots). Following Pb
exposure, toxicological studies have demonstrated that Pb induces markers of systemic inflammation in
blood (Fernandez-Cabezudo et al.. 2007; Iavicoli et al.. 2006; Chen et al.. 2004; Dvatlov and Lawrence.
2002; Miller etal.. 1998; Heo et al.. 1997; Heo et al.. 1996). as well as increases in C-reactive protein in
cardiac tissue (Roshan et al.. 2011). In addition, Pb was also found to induce interleukin (IL)-8, which
mediated vessel intima hyperplasia in human endothelial cells (Zeller et al.. 2010). Similarly, Pb exposure
in rats increased aortic media thickness, media-lumen ratio, and medial collagen content (Zhang et al..
2009). Exposure to Pb also promoted thrombus formation in rats (Shin et al.. 2007). as well as cellular
perforations and membrane blebbing in endothelial cells (van Strijp et al.. 2023). In agreement with these
toxicological studies demonstrating Pb-induced inflammation, an epidemiologic study found that higher
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BLLs in children were correlated with higher serum levels of IL-4 (Lutz et al.. 1999). which can stimulate
the liver to produce additional coagulation factors and further the pro-atherosclerotic environment.
Moreover, as discussed in the metabolic effects appendix (Section 9.2), exposure to Pb can also result in
the upregulation of cholesterol, another key contributor to developing atherosclerosis. Taken together,
evidence of a pro-atherosclerotic environment is important given that atherosclerosis can lead to plaque
and thrombosis (i.e., clot) formation. If dislodged, those plaques could obstruct blood flow to the heart or
stimulate intravascular clotting (Karolv et al.. 2007). both of which could result in acute myocardial
ischemia. If the dislodged plaque obstructs blood flow to the brain, there is potential for stroke.
Impaired vascular function (e.g., resulting from Pb-induced oxidative stress) can also lead to
increases in BP through vasoconstriction. Increases in BP may then exacerbate IHD or heart failure
through conduction abnormalities or arrythmias, and further impair vascular function. For example, in
patients with high BP, changes in arterial shear stress due to changes in blood flow (i.e., laminar versus
turbulent) are associated with impaired vascular function (Khder et al.. 1998). which, as noted above,
could lead to a worsening of IHD or heart failure. Importantly, there are numerous studies demonstrating
that Pb can increase measures of BP: (Nunes et al.. 2015; Xu et al.. 2015; Fioresi et al.. 2014; Mohammad
et al.. 2010; Zhang et al.. 2009; Badavi et al.. 2008; Grizzo and Cordellini. 2008: Reza et al.. 2008; Bravo
et al.. 2007; Robles et al.. 2007; Hevdari et al.. 2006; Bagchi and Preuss. 2005; Nakhoul et al.. 1992). In
addition, a toxicological study has demonstrated that Pb exposure can result in conduction abnormalities
and potential arrythmia (Reza et al.. 2008).
The second pathway by which Pb exposure may result in the cardiovascular associations reported
in epidemiologic studies is through activation of the RAAS, which is responsible for fluid homeostasis
and BP regulation. Exposure of experimental animals to Pb increases ACE activity; plasma kininase II;
kininase I; and kallikrein activities in plasma, aorta, heart, and kidney, as well as renal angiotensin II
positive cells (Rodriguez-Iturbe et al.. 2005; Sharifi et al.. 2004; Carmignani et al.. 1999). These changes
can result in increases in BP, which as noted above, could lead to worsening of IHD or heart failure
potentially through conduction abnormalities, arrythmia, and/or impaired vascular function. Additional
information on the effect of Pb on the RAAS system is discussed in the renal effects appendix. This
summary includes a discussion of Pb accumulation in the kidney resulting in cellular damage, thereby
increasing the potential for RAAS disfunction (see Appendix 5).
The third pathway by which Pb exposure may result in the cardiovascular associations reported in
epidemiologic studies is through modulation of the ANS. As noted in the nervous system appendix, Pb
can deposit in the brain where it causes cellular damage and altered neurological function (see
Appendix 3). Similarly, it has also been shown that exposure to Pb can modulate autonomic tone
(e.g., increased sympathetic tone) to the heart and vasculature, possibly through stimulation of the P2X4
and P2X7 receptors in satellite glial cells (Zhu et al.. 2019; Zhu et al.. 2018). Shifts toward increased
sympathetic nervous system tone may result in increases in heart rate and BP as well as decreases in
vascular function, which as mentioned above, could exacerbate IHD and/or heart failure. It is therefore
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important to note evidence from animal toxicological studies for increases in heart rate (Simoes et al..
2011; Badavi et al.. 2008; Lai et al.. 2002) and changes in HRV consistent with a shift toward increased
sympathetic tone (Zhu et al.. 2019; Shvachiv et al.. 2018; Zhu et al.. 2018; Geraldes et al.. 2016)
following Pb exposure. Similarly, evidence from an animal toxicological study suggests that Pb exposure
can result in conduction abnormalities or arrhythmia (Reza et al.. 2008). Conduction abnormalities or
arrhythmia could exacerbate IHD and/or HF. Taken together, there are potential pathways by which ANS
modulation may lead to worsening of IHD or HF, thereby increasing the risk for mortality.
When considering the available evidence presented throughout this appendix, there are plausible
pathways connecting Pb exposure to the cardiovascular associations reported in epidemiologic studies
(Figure 4-26). Thus, these proposed pathways provide biological plausibility for the associations reported
in epidemiologic studies between Pb and IHD, MI, stroke, and therefore, mortality.
4.12 Summary and Causality Determination
A large body of health evidence published since the 2013 Pb ISA continues to demonstrate a
causal relationship between exposure to Pb and cardiovascular health effects. The 2013 Pb ISA concluded
that the evidence supported a causal relationship between exposure to Pb and hypertension and increased
BP in adults, as well as between Pb exposure and CHD (based largely on epidemiologic studies of CVD-
related mortality). For other cardiovascular-related outcomes, the evidence was suggestive of but not
sufficient to infer a causal relationship for subclinical atherosclerosis and inadequate to infer the presence
or absence of a relationship with cerebrovascular disease (Table 4-2). More recent studies greatly expand
the evidence base discussed in the 2013 Pb ISA and serve to strengthen the support for relationships
between exposure to Pb and a number of cardiovascular-related health effects. In particular, there is
substantially more evidence of hypertension, increases in BP, and cardiovascular-related mortality
following exposure to Pb. Moreover, there is additional health evidence for effects such as changes in
cardiac electrophysiology (e.g., ECG measures of cardiac depolarization, repolarization, and HRV),
arrythmia, and markers of atherosclerosis. Thus, in the current ISA, the evidence supports a causal
relationship between exposure to Pb and cardiovascular effects and cardiovascular-related mortality.6
After a brief discussion of the health evidence and key uncertainties found in the 2013 Pb ISA, the rest of
this summary and causal determination section discusses the health evidence and rationale for the causal
determination reached in this Pb ISA. This discussion will rely upon the framework for causality
determinations described in the preamble to the ISAs (U.S. EPA. 2015). Key health evidence supporting
this determination is also summarized in Table 4-2.
6The current ISA follows the approach of more recent ISAs, including the 2019 Particulate Matter and 2020 Ozone
ISAs, in making a single causality determination for cardiovascular effects. Additional information regarding this
decision can be found in Section 4.1 of this appendix.
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In the 2013 Pb ISA, the strongest evidence for an effect of Pb on cardiovascular outcomes was on
BP and CVD-related mortality. Prospective epidemiologic studies clearly supported the relationship
between biomarkers of Pb exposure and hypertension incidence and changes in BP (U.S. EPA, 2013).
The prospective evidence was supported by meta-analyses that underscored the consistency and
reproducibility of Pb-associated increases in BP and hypertension and epidemiologic studies that adjusted
for a wide range of potential confounders to reduce uncertainty due to potential unmeasured confounding
(U.S. EPA, 2013). With respect to cardiovascular-related mortality, the previous Pb ISA (U.S. EPA,
2013) described longitudinal studies in adult cohorts in a number of locations reporting that biomarkers
of Pb exposure were associated with risk of mortality from MI, IHD, or CHD, with the strongest of these
associations being with MI mortality. In addition, epidemiologic studies reviewed in the 2013 Pb ISA
included some evidence of a positive association between exposure to Pb and changes in cardio
electrophysiology (e.g., changes in HRV and QT interval) and atherosclerotic plaque formation. Key
uncertainties noted with respect to the epidemiologic evidence from the last review included inconsistent
evidence for BP changes in children and uncertainty in the level, timing, frequency, and duration of Pb
exposure contributing to the reported cardiovascular effects in adults. That is, given the appreciable
history of exposure in decades past (see Appendix 2, Section 2.4.1), and that Pb accumulates in the body
over a lifetime, the extent to which past Pb exposures contribute to the BLLs and positive associations
reported in epidemiologic studies remains uncertain.
In the 2013 Pb ISA, animal toxicological studies provided supporting evidence and biological
plausibility for the associations observed in epidemiologic studies, particularly with respect to BLLs and
changes in BP and/or hypertension. Increases in BP following exposure to Pb were generally reported in
animal toxicological studies. The previous ISA further noted toxicological studies indicating the
production of oxidative stress species that could inactivate the vasodilator, nitric oxide, which could
potentially lead to increased vasoconstriction, and thus, increases in BP. Animal toxicological studies
discussed in the last review also provided at least some evidence that exposure to Pb may contribute to a
pro-atherosclerotic environment and result in changes in HRV.
More recent studies greatly expand the evidence base from the 2013 Pb ISA and serve to
strengthen support for the relationship between exposure to Pb and cardiovascular effects in adults. In
particular, the strongest evidence continues to be Pb's effect on increases in BP. Numerous additional
epidemiologic studies published since the last review report positive associations between measures of Pb
in the body and increases in BP. Nationally representative cross-sectional studies in countries including
the United States and Canada reported positive associations between increases in BLLs and changes in
SBP, DBP, or both (Huang, 2022; Qu et al., 2022; Everson et al., 2021; Teye et al., 2020; Obeng-Gyasi et
al„ 2018; Lee et al., 2016a; Hara et al„ 2015; Bushnik et al„ 2014; Hicken et al„ 2013; Zota et al„ 2013b;
Scinicariello et al., 2011). These nationally representative studies of adult cohorts (with most participants
born before 1970, some in the 1930s) generally reported positive increases in BP (mmHg) with mean
BLLs -1.5-3 (ig/dL. Consistent with these nationally representative studies, smaller cross-sectional
studies generally reported positive associations between measures of Pb burden and changes in BP (Yan
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et al., 2022; Chung et al., 2020; Wang et al., 2020; Guo et al., 2019; Gambelunghe et al.. 2016; Ettinger et
al., 2014) within a slightly larger range of mean BLLs within each study (~1.5- 8.5 (ig/dL). While not all
studies reported positive associations with SBP or DBP (e.g., KYu et al.. 2020; Ettinger et al.. 2014)1). the
generally positive cross-sectional results were consistent with a longitudinal analysis in a small
Bangladeshi cohort. This study indicated there was an annual increase in SBP associated with the largest
quartile of baseline BLLs compared with the lowest quartile Bulka et al. (2019). The majority of recent
analyses consider a wide range of confounders including demographics, comorbid conditions,
antihypertensive medication use, and other co-exposures to metals such as Cd. In addition, an extensive
amount of literature also considered effect measure modifiers, including, sex, age, and race, among
others. Combined with epidemiologic results from the previous ISA and AQCDs, there is clear and
substantial evidence that increasing body Pb levels is associated with increases in measures of BP.
However, uncertainty remains regarding the role of extensive historical exposure (magnitude, duration,
timing) of these cohorts.
The epidemiologic associations summarized above are coherent with animal toxicological studies
published since the 2013 Pb ISA that examined BP. BLLs in these animal studies were <30 (ig/dL, and
most of these studies reported that in animals exposed to Pb, there were increases in BP when compared
with control treated animals (Zhu et al„ 2019; Shvachiy et al„ 2018; Zhu et al„ 2018; Nunes et al., 2015;
Silva et al., 2015; Xu et al., 2015; Fioresi et al., 2014; Gaspar and Cordellini, 2014). It should be noted
that although these studies found some measure of BP at some time point to be increased following
exposure to Pb, there was variability among studies with respect to which measure of BP increased
(e.g., SBP or DBP) and the timing of those increases. Moreover, there was a single animal toxicological
study in rats that reported no changes in measures of BP following a Pb drinking water exposure
(Wildemann et al., 2015). Nonetheless, when considered in total, these animal toxicological studies
provide clear evidence for exposure to Pb resulting in increases in measures of BP. These animal
toxicological studies are coherent with, and provide support for, the mostly positive associations reported
in epidemiologic studies between body Pb levels and BP increases.
As noted above, a number of prospective cohort studies evaluated in the 2013 Pb ISA (U.S. EPA,
2013) and in the 2006 Pb AQCD (U.S. EPA, 2006) indicated positive associations between Pb
biomarkers of exposure and cardiovascular mortality. Moreover, the results of these previously reviewed
studies remained positive when controlling for a wide range of potential confounders. Since the
publication of the 2013 Pb ISA, additional evidence of cardiovascular-related mortality has been reported.
In an analysis of the NHANES III cohort, a 1 (ig/dL increase in BLLs was associated with HRs of 1.10
(95% CI: 1.05, 1.15) for CVD mortality and 1.14 (95% CI: 1.08, 1.20) for IHD mortality (Lanphear et al„
2018). Consistent with these results, additional studies analyzing NHANES cycles reported associations
of similar magnitudes between BLLs and CVD-related mortality (Duan et al„ 2020; Ruiz-Hernandez et
al., 2017; Aoki et al., 2016; van Bemmel et al„ 2011). These more recent studies also reported that
associations between BLLs and CVD-related mortality remained positive after accounting for risk factors
such as physical activity, serum cholesterol, (Lanphear et al„ 2018; Ruiz-Hernandez et al„ 2017) and Cd
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levels in blood or urine (Aoki et al.. 2016) (Table 4-16). In addition, Duan et al. (2020) specifically
evaluated NHANES participants enrolled in cycles between 1999 and 2014 (mortality data included
through 2015), with ~7 years of mortality follow-up. Although some members of this population may
have had lower Pb exposures due to the phaseout of leaded gasoline, especially when compared with
studies assessing adults in NHANES II (1976-1980) and NHANES III (1988-1994), the vast majority of
the participants were born well before the phaseout.
Epidemiologic studies of mortality are consistent not only with the large amount of evidence for
changes in BP and hypertension described above, but also with evidence of associations between blood or
bone Pb levels and other cardiovascular outcomes. Studies using the NAS cohort of older adult men
indicated an association between patella Pb levels and incident IHD (Ding et al.. 2019; Ding et al.. 2016).
Additionally, a series of 10-year CVD risk evaluations using KNHANES data observed increased 10-year
CVD risk with increasing BLLs (Nguyen et al.. 2021; Park and Han. 2021; Choi et al.. 2020; Cho et al..
2016). These studies are also consistent with a series of NAS analyses presented in the 2013 Pb ISA
indicating an association between bone Pb levels and a prolonged QT interval (Eum et al.. 2011; Park et
al.. 2009; Cheng et al.. 1998). Recent studies among children are less consistent than those in adults.
However, there is some evidence to support clinically relevant changes in HRV (Halabickv et al.. 2022)
and increases in SBP and TPR (Gump et al.. 2011) following an acute stressor.
Toxicological studies evaluated in the 2013 Pb ISA demonstrated increased incidence of
arrhythmia, atrioventricular block, and a prolonged ST segment interval in Pb-exposed animals (Reza et
al.. 2008). That said, an additional toxicological study published since the last review reported no change
in the PR, QRS, or QT segments in Pb-exposed rats (Wildemann et al.. 2015). Similarly, although more
limited and/or mixed, there is at least some epidemiologic and animal toxicological evidence for changes
in heart rate and HRV (Section 4.7) and potential indicators of atherosclerosis following exposure to Pb.
Notably, a toxicological study published since the 2013 Pb ISA reported a statistically significant increase
in the diameter of the cells of the aorta, as well as changes in the shape (i.e., loss of curvature) of the
aortic internal elastic lumen in Pb-exposed rats, relative to control rats. This study also reported a
statistically significant increase in proliferating cell nuclear antigen in rat cardiac tissue in Pb-exposed rats
(p < 0.05), potentially consistent with the type of cellular proliferation that is involved in atherosclerotic
plaque growth (Xu et al.. 2015). Moreover, these results are consistent with a study discussed in the 2013
Pb ISA demonstrating increased aortic media thickness, media-lumen ratio, and medial collagen content
following exposure to Pb (Zhang et al.. 2009).
In support of epidemiologic studies reporting positive associations between BLLs and CVD-
related mortality, animal and in vitro toxicological evidence provides plausible pathways by which
exposure to Pb could lead to serious CVD-related outcomes such as IHD, MI, and/or stroke. In brief, one
such pathway posits that exposure to Pb resulting in oxidative stress and systemic inflammation could
potentially lead to impaired vascular function, a pro-atherosclerotic environment, and increases in BP.
Importantly, there is animal toxicological evidence demonstrating all these effects following exposure to
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Pb (Section 4.8). In addition, these effects, in particular atherosclerosis and increases in BP, can set the
stage for an MI or stroke that could result in mortality. More information on this and other potential
pathways can be found in Section 4.11, in which each potential pathway is described in detail. Several
recent epidemiologic studies have been published further supporting this association. Many of the
epidemiologic studies evaluated in this appendix utilize data from large population-based health surveys
(e.g., NHANES). Due to the higher prevalence of cardiovascular disease in older populations, most
available studies examine populations born before the phaseout of leaded gasoline. Although some
members of these study populations may have had lower Pb exposures due to the phaseout of leaded
gasoline, especially when compared with studies assessing adults who participated in older studies, such
as NHANES II (1976-1980) and NHANES III (1988-1994), the vast majority of participants across older
and more recent studies were born well before the phaseout. It is also important to note that Pb in blood at
a particular time point may be reflective of more recent exposures, or due to mobilization of Pb from bone
stores. While uncertainty remains regarding the role of extensive historical exposure (magnitude,
duration, timing) in health outcomes assessed in these cohorts, there is still sufficient evidence that
supports the use of blood Pb as a valid and reliable biomarker for assessing associations of Pb with long-
term effects, such as cardiovascular disease and mortality (Ruiz-Hernandez et al.. 2017).
The collective evidence is sufficient to conclude that there is a causal relationship between
Pb exposure and cardiovascular effects and cardiovascular-related mortality. Evidence from
epidemiologic studies indicates consistent associations between Pb biomarkers and cardiovascular
endpoints such as BP (Section 4.3.1.1), hypertension (Section 4.3.1.2), and mortality (Section 4.10). This
evidence was further supported by experimental animal studies (Section 4.3.2, Section 4.11). Studies
relying on bone biomarkers to assess Pb exposure provided consistent evidence for an association
between cumulative exposures and chronic health outcomes, such as hypertension and premature
mortality. Evidence of this effect was further supported by cross-sectional studies primarily evaluating
concurrent BLL levels and cardiovascular health effects. It is also important to note that Pb in blood at a
particular time point may be reflective of more recent exposures, or due to mobilization of Pb from bone
stores. While uncertainty remains regarding the role of extensive historical exposure (magnitude,
duration, timing) in health outcomes assessed in these cohorts, there is still sufficient evidence that
supports the use of blood Pb as a valid and reliable biomarker for assessing associations of Pb with long-
term effects, like mortality (Ruiz-Hernandez et al.. 2017). While much of the evidence between Pb
biomarkers and cardiovascular effects is consistent, some specific cardiovascular outcomes are examined
in relatively few studies, and the results across these studies are inconsistent (see Sections 4.7-4.9).
Furthermore, uncertainties remain regarding the timing, frequency, and duration of the Pb exposures that
contribute to cardiovascular health effects. Yet, even after the consideration of these uncertainties, the
overall evidence base strongly indicates that exposure to Pb is associated with numerous cardiovascular
effects including increases in BP and cardiovascular-related mortality.
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Table 4-2 Summary of evidence indicating a causal relationship between Pb
exposure and cardiovascular effects and cardiovascular-related
mortality
Rationale for
Causality Key Evidence* References* Associatedwfth
Determination3 Associated witn tnects
Generally consistent
evidence from
epidemiologic studies
of BP in adults
Epidemiologic studies
consistently demonstrating
increases in at least some
measure of BP and Pb
biomarkers
Hara et al. (2015)
Hicken et al. (2012)
Hicken et al. (2013)
Obenq-Gvasi et al. (2018)
Scinicariello et al. (2011)
Teve et al. (2020)
Zota et al. (2013b)
Everson et al. (2021)
Huang (2022)
Obenq-Gvasi (2019)
Tsoietal. (2021)
Lee et al. (2016a)
Bushnik et al. (2014)
Qu et al. (2022)
Lopes et al. (2017a)
Chung et al. (2020)
Yan et al. (2022)
Gambelunqhe et al. (2016)
Bulka et al. (2019)
Mean blood Pb: -1.0 to
3 |jg/dL
Generally consistent
evidence from
epidemiologic studies
of hypertension
Epidemiologic studies
consistently demonstrating
increases in incident
hypertension risk with Pb
biomarkers
Mostly positive associations
between prevalent
hypertension and Pb
biomarkers
Gambelunqhe et al. (2016)
Zheutlin et al. (2018)
Huang (2022)
Tsoietal. (2021)
Miao et al. (2020)
Scinicariello et al. (2011)
Lee et al. (2016b)
Lee et al. (2016a)
Choi et al. (2018)
Qu et al. (2022)
Lopes et al. (2017a)
Mean blood Pb: ~2.5-
5 |jg/dL
Mean bone Pb: -20 (tibia) ¦
27 (patella) |jg/g
Mean blood Pb: ~1.5-
3.5 |jg/dL
Generally consistent
evidence from
epidemiologic studies
of cardiovascular
mortality
Ruiz-Hernandez et al. (2017)
Duan et al. (2020)
Aoki et al. (2016)
Obenq-Gvasi et al. (2021)
Lin et al. (2011)
Epidemiologic studies
consistently demonstrating
increases in cardiovascular
mortality risk with Pb
biomarkers
Menke etal. (2006)
Lanphear et al. (2018)
van Bemmel et al. (2011)
Cook et al. (2022)
Mean blood Pb: ~1.5-
3.2 |jg/dL
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Rationale for
Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels
Associated with Effects0
Generally consistent Mostly positive associations Jain et al. (2007)
evidence from
epidemiologic studies
of ischemic heart
disease
between incident IHD and Pb
biomarkers
Mostly positive associations
between estimates of 10-yr
CHD risk and Pb biomarkers
Ding et al. (2016)
Ding et al. (2019)
Cho et al. (2016)
Choi et al. (2020)
Park and Han (2021)
Nguyen et al. (2021)
Mean blood Pb: -6.5 |jg/dL
Mean bone (patella) Pb:
~30 |jg/g
Mean bone (tibia) Pb:
~23 |jg/
Mean blood Pb: ~3 |jg/dL
Generally consistent
evidence from
epidemiologic studies
of cardiac function
Mostly positive associations
between left ventricle
structure/function and Pb
biomarkers
Yang etal. (2017)
Lind et al. (2012)
Chen etal. (2021)
Mean blood Pb: ~2-5 |jg/dL
Limited evidence
from epidemiologic
studies for changes
in HRV
A single study reported a
change in HRV following a
stress response in children
Halabickv et al. (2022)
Mean blood Pb: ~3-6 |jg/dL
Limited evidence
from epidemiologic
studies for
atherosclerosis
A small number of studies Wan et al. (2021)
demonstrated development of Kim et al (2021)
atherosclerosis within
different populations Qin etal. (2021)
Mean blood Pb: 1.5-3 |jg/dL
Consistent evidence
from animal
toxicological studies
of BP
Animal toxicological studies
consistently demonstrating
increases in at least some
measure of BP
Fioresi et al. (2014)
Nunes et al. (2015)
Xu etal. (2015)
Silva et al. (2015)
Shvachiv et al. (2018)
Gasparand Cordellini (2014)
Zhu etal. (2018)
Zhu etal. (2019)
Mean blood Pb: -8-30 |jg/dL
Limited evidence
from animal
toxicological studies
for changes in HRV
A small number of studies
demonstrated changes in at
least some measure of HRV
(e.g., LF)
Shvachiv et al. (2018)
Zhu etal. (2018)
Zhu etal. (2019)
Mean blood Pb: ~24-
28 |jg/dL
4-78
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Rationale for
Causality Key Evidence* References* Associatedwfth
Determination3 Associated witn tnects
Limited but
consistent evidence
from animal
toxicological studies
for structural
changes consistent
with the development
of atherosclerosis
A single animal toxicological
study reported an increase in
the aortic media thickness,
media-lumen ratio, and
medial collagen content
following Pb exposure
A single animal toxicological
study reporting an increase in
the diameter of the cells of
the aorta, changes in the
shape of the aortic internal
elastic lumen, and an
increase in proliferating cell
nuclear antigen in rat cardiac
tissue
Zhang et al. (2009)
Mean blood Pb: -28 |jg/dL
Xu et al. (2015)
Mean blood Pb: ~20-
25 |jg/dL
Biological Plausibility A few well-defined potential Section 4.11 NA
pathways by which exposure
to Pb could reasonably result
in the health outcomes
reported in epidemiologic
studies
BP = blood pressure; CHD = coronary heart disease; HRV = heart rate variability; IHD = ischemic heart disease; LF = low
frequency; NA = not available; Pb = lead; yr = year(s).
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the
Preamble to the ISAs (U.S. EPA. 2015).
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and,
where applicable, to uncertainties or inconsistencies. References to earlier sections indicate where the full body of evidence is
described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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4.13 Evidence Inventories - Data Tables to Summarize Study Details
Table 4-3 Epidemiologic studies of Pb exposure and blood pressure
Reference and Study
Study Design Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Cross-Sectional Studies
Blood Pb (ICP-MS) (|jg/dL) BP (SBP, DBP, Linear models adjusted for BP change (mmHg) per doubling of
Hara et al. (2015) NHANES
United States
n = 12,725
>20 yr
NHANES 2003-
2010
Cross-sectional
Average
individual born
-1957
GM (IQR):
See Figure 4-4
Age at measurement:
Mean (SD)
Black Women: 48.31 (6.8)
Hispanic Women: 48.1 (16.8)
White Women: 53.0 (8.4)
Black Men: 47.7 (16.9)
Hispanic Men: 46.1 (6.8)
White Men: 53.1 (18.6)
PP, MAP) ethnicity, sex, age, BMI,
heart rate, hematocrit, serum
total calcium y-
glutamyltransferase, cotinine,
dietary sodium to potassium
intake ratio, college
education, antihypertensive
drug treatment
blood Pb
See Figure 4-4b
Hicken et al.
(2012)
United States
NHANES 2005-
2008
Cross-sectional
NHANES
n =10,971
>20 yr
Average
individual born
-1963
Blood Pb (ICP-MS) (|jg/dL)
Mean (Median)
See Figure 4-6
Age at measurement
Mean (SD)
White Men: 45.6 (15.8)
Black Men: 40.6 (14.4)
White Women: 47.3 (16.7)
Black Women: 42.4 (15.1)
BP (SBP, DBP,
PP)
Linear regression adjusted
forage, BMI, heavy alcohol
use, smoking status,
diabetes diagnosis,
antihypertensive medication
use, and dietary intake of
sodium, calcium, and
potassium
Change in BP (mmHg)
See Figure 4-6b
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Hicken et al.
(2013)
United States
NHANES 2005-
2008
Cross-sectional
NHANES
n = 4,470
Nonpregnant
adults (>20 yr)
Average
individual born
-1962
Blood Pb (ICP-MS) (pg/dL)
Mean (SD)
lack: 1.9 (2.2)
White: 1.7 (0.9)
Age at measurement:
Mean (SD)
Black: 42.2 (16.4)
White: 47.1 (10.8)
BP (SBP, DBP)
Linear regression adjusted
for race/ethnicity, age, sex,
high school education, family
poverty, hematocrit, BMI,
heavy alcohol use, smoking
status, and diabetes
BP (mmHg) per doubling of blood Pbb
SBP
Black 3.2 (1.5, 5.0)
High Depression 5.6 (2.0, 9.2)
Low Depression 1.8 (0.2, 3.5)
White 1.0 (-0.3, 2.4)
High Depression 1.2 (-0.5, 2.9)
Low Depression 1.0 (-0.6, 2.6)
DBP
Black 1.8 (0.7, 2.8)
White 0.9 (0.1, 1.8)
Obenq-Gvasi et NHANES
al. (2018)
United States
2007-2010
Cross-sectional
n = 12,153
>20 yr
Average
individual born
-1958
Blood Pb (ICP-MS) (pg/dL) BP (SBP, DBP)
Mean (SD)
Q1 (0-2): 1.09 (0.01)
Q2 (2-5): 2.78 (0.02)
Q3 (5-10): 6.40 (0.10)
Q4 (>10): 16.11 (1.40)
Age at measurement
Mean (SD):
Q1
Q2
Q3
Q4
44.25 (0.32)
56.05 (0.54)
54.77 (1.13)
47.56 (2.56)
Linear regression adjusted
for age, sex, race/ethnicity,
BMI, antihypertensive
medication
BP (mmHg) and In-blood Pbbc
DBP 0.268 (0.079, 0.458)
SBP 0.052 (-0.233, 0.458)
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Scinicariello et al.
(2010)
United States
NHANES III
1988-1994
Cross-sectional
NHANES III
n =6,016
>17 yr
Average
individual born
-1963, -1949,
and in or
before -1931
Blood Pb (GFAAS) (|jg/dL)
Mean (SE)
Overall, 2.99 (0.09)
NH White 2.87 (0.09)
NH Black 3.59 (0.20)
Mexican American 3.33 (0.11)
Age at measurement:
17-3947%
40-59 30.9%
>60 22.1%
BP (SBP, DBP)
Multivariable linear
regression adjusted for age,
sex, education, smoking
status, alcohol intake, BMI,
serum creatinine levels,
serum calcium, glycosylated
hemoglobin, and hematocrit
BP (mmHg) and blood Pb
SBP
NH White 0.707 (0.216, 1.199)
NH Black 1.615 (1.007, 2.223)
Mexican American 0.471 (0.062, 0.879)
DBP
NH White -0.094 (-0.741, 0.553)
NH Black 1.261 (0.716, 1.805)
Mexican American 0.414 (-0.001, 0.83)
Significant interactions with blood Pb and
ALAD genotype observed in relation to
SBP for NH white and NH Black
individuals
Scinicariello et al.
(2011)
United States
NHANES 1999-
2006
Cross-sectional
NHANES
n =16,222
>20 yr with
blood Pb
<10 |jg/dL
Average
individual born
-1959
Blood Pb (ICP-MS) (ug/dL)
Mean (SE)
See Figure 4-5
Age at measurement:
Mean (SE)
White men: 47.14 (0.37)
White women: 49.64 (0.36)
Black men: 42.86 (0.37)
Black women: 45.10 (0.42)
Mexican-American men: 37.64
(0.48)
Mexican-American women:
40.67 (0.65)
BP (SBP, DBP,
PP)
Multivariable logistic and
linear regression models
adjusted for age, BMI, self-
reported diabetes alcohol
ingestion, smoking status,
education, serum
creatinine, serum total
calcium, sodium, hematocrit,
and blood Cd
BP (mmHg) and twofold increase in blood
Pbb
See Figure 4-5
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Teve et al. (2020) NHANES
n = 30,467
United States 20-79 yr
NHANES 1999-
2016
Cross-sectional
Average
individual born
-1965
Blood Pb (ICP-MS)d (|jg/dL)
Median (IQR)
NH White
Men: 1.50 (0.99, 2.29)
Women: 1.06 (0.69, 1.60)
NH Black
Men: 1.60 (1.00, 2.60)
Women: 1.11 (0.71, 1.77)
Hispanic
Men: 1.58 (0.99, 2.43)
Women: 0.95 (0.62, 1.51)
Other race
Men: 1.54 (1.05, 2.39)
Women: 1.16 (0.75, 1.79)
BP (SBP, DBP)
Linear regression adjusted
for age/ethnicity, age,
gender, education level, BMI,
and PIR
BP (mmHg)e
SBP
NH White: 0.34 (0.11, 0.57)
NH Black: 0.67 (0.29, 1.05)
Hispanic: 0.10 (-0.01, 0.21)
Other: 0.44 (-0.51, 1.39)
DBP
NH White: 0.38 (0.19, 0.57)
NH Black: 0.36 (0.06, 0.66)
Hispanic: -0.08 (-0.21, 0.05)
Other: 0.27 (-0.15, 0.69)
Age at measurement
Mean age
NH white men: 46.37
NH white women: 47.00
NH Black men: 43.09
NH Black women: 43.28
Hispanic men: 39.67
Hispanic women: 40.51
Other men: 42.92
Other women: 43.54
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Zota et al.
(2013b)
United States
1999-2008
Cross-sectional
NHANES
n =8,194
40-65 yr
Average
individual born
-1953
Blood Pb (ICP-MS) (pg/dL)
GM 1.69
Geometric SE (GSE) 0.02
Quintiles GM (GSE)
BP (SBP, DBP) Logistic and linear regression OR(Q5vs. Q1)
Q1
Q2
Q3
Q4
Q5
0.76 (0.01)
1.25 (0.00)
1.67 (0.00)
2.25 (0.01)
3.88 (0.03)
Age at measurement
Mean: 50.9
SE: 0.15
adjusted for age, educational
attainment, race/ethnicity,
smoking, alcohol
consumption, marital status,
and antihypertensive
medication use
Elevated SBP (>140 mmHg)
All participants: 1.23 (0.92, 1.65)
LowAL: 1.14 (0.79, 1.66)
High AL: 1.40 (0.99, 1.97)
Elevated DBP (>90 mmHg)
All participants: 1.77 (1.25, 2.50)
LowAL: 1.46 (0.80, 2.68)
High AL: 2.28 (1.33, 3.91)
BP change (mmHg, Q5 vs. Q1)
SBP
All participants: 0.36 (-1.07, 2.33)
LowAL: 0.67 (-1.24, 2.58)
High AL: 1.60 (-0.62, 3.82)
DBP
All participants: 1.76 (0.75, 2.78)
LowAL: 1.72 (0.62, 2.95)
High AL: 2.01 (0.24, 3.79)
Everson et al. NHANES Blood Pb (ICP-MS) (pg/dL) BP (SBP, DBP) Linear regression models BP (mmHg)
(2021) n =2,413 Median: 1.5 adjusted for age, age2, race, SBP 0.73 (0.03, 1.44)
sex, BMI, and smoking status „ „„ , „ „„
a DBP 0.41 (-0.10, 0.92)
United States 20-59 yr Age at measurement:
Range 20-59 yr
NHANES 1999- Average
2004 individual born
-1962
Cross-sectional
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Huang (2022)
United States
NHANES
n = 32,289
>20 yr
NHANES 1999- Average
2018 individual born
-1958
Cross-sectional
Blood Pb (ICP-MS) (pg/dL)
Mean (SD) 1.73 (1.71)
Age at measurement
Mean (SD) 49.68 (18.04)
BP (SBP, DBP)
Linear regression models
adjusted for age, sex, race,
education, family income
poverty ratio, BMI, alcohol
use, and smoking
BP (mmHg)
SBP
All 0.30 (0.19, 0.42)
Men
Mexican American 0.01 (-0.13, 0.34)
Other Hispanic 0.07 (-0.31, 0.45)
NH White 0.44 (0.22, 0.66)
NH Black 0.37 (0.07, 0.67)
Other Race 0.49 (-0.04, 1.03)
Women
Mexican American 0.14 (-0.28, 0.57)
Other Hispanic 0.84 (-0.15, 1.83)
NH White 0.63 (0.22, 1.04)
NH Black 0.99 (0.48, 1.50)
Other Race 0.49 (-0.35, 1.34)
DBP
All 0.23 (0.14, 0.32)
Men
Mexican American 0.08 (-0.11, 0.26)
Other Hispanic -0.20 (-0.51, 0.11)
NH White 0.40 (0.22, 0.58)
NH Black 0.26 (0.00, 0.51)
Other Race 0.05 (-0.37, 0.48)
Women
Mexican American 0.08 (-0.25, 0.40)
Other Hispanic 0.42 (-0.30, 1.14)
NH White 0.74 (0.41, 1.07)
NH Black 0.80 (0.40, 1.20)
Other Race 0.16 (-0.47, 0.79)
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Obenq-Gvasi
(2019)
United States
NHANES 2009-
2016
Cross-sectional
NHANES
young adults
(18-44 yr)
(n = 7,730),
middle-aged
adults (45-
65 yr)
(n = 5,744)
Average
individual born
-1981 and
-1957
Blood Pb (ICP-MS) (|jg/dL)
mean (SE):
young adults:
1.03 (0.026)
Middle-aged adults:
1.62 (0.044)
BP (SBP, DBP)
Logistic regression adjusted
for sex, BMI, income,
ethnicity, alcohol
consumption, and smoking
OR (above/below 5 |jg/dL)b
SBP >120 mmHg
Young adults: 1.21 (1.07, 1.38)
Middle-aged adults: 1.32 (1.14, 1.52)
DBP >80 mmHg
Young adults: 1.32 (1.10, 1.58)
Middle-aged adults: 1.16 (0.98, 1.38)
Tsoietal. (2021) NHANES
United States
NHANES 1999-
2016
Cross-sectional
N = 39,477
adults >20
Average
individual born
-1960
Blood Pb ICP-MS (|jg/dL)
Median 1.30
Q1 <0.89
Q2 0.89-1.30
Q3 1.30-2.10
Q4 >2.10
Age at measurement:
Mean (SE)
Hypertensive
54.08 (0.23) yr
Non-hypertensive
39.87 (0.19) yr
BP (SBP) Multivariable linear
regression adjusted for age,
sex ethnicity, waist
circumference, PIR,
education, ever cigarette
smoking, diabetes, and stage
3-5 chronic kidney diseases
SBP (mmHg)
For every doubling of blood Pbb
0.52 (0.19, 0.86)
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Lee etal. (2016a)
South Korea
KNHANES IV
(2008-2009), V
(2010-2012), and
VI (2013)
Cross-sectional
Korean
NHANES
n = 11,797
>19 yr
Average
individual born
in or before
-1991
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
GM (95% CI)
Male: 2.396 (2.362, 2.430)
Female: 1.919 (1.889, 1.949)
Age at measurement >19 yr
BP (SBP, DBP)
Linear models adjusted for
sex, age, residence area,
education level, smoking,
drinking status, BMI, physical
activity, serum creatinine,
and hemoglobin
BP (mmHg) doubling of blood Pbb
SBP
All 0.73 (0.09, 1.36)
Male: 0.30 (-0.53, 1.14)
Female: 1.08 (0.26, 1.90)
DBP:
All 0.71 (0.29, 1.13)
Male: 0.59 (0.01, 1.17)
Female: 0.80 (0.28, 1.33)
Bushnik et al.
(2014)
Canada
2007-2011
Cross-sectional
Canadian
Health
Measures
Survey
n = 4,550
Nonpregnant
individuals
aged 40-79
Blood Pb (ICP-MS) (pg/dL)
Mean 1.64 (1.58-1.71)
Age at measurement:
mean: 55.4
BP (SBP, DBP)
Linear regression adjusted
for age, sex, education,
smoking, alcohol, physical
activity, BMI, non-HDL
cholesterol, diabetes, chronic
kidney disease, family history
of high BP, antihypertension
medication use
BP (mmHg)
SBP
See Figure 4-11
DBP
See Figure 4-12
Average
individual born
-1954
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Qu et al. (2022)
China
2017-2018
Cross-sectional
China National
Human
Biomonitoring
Study
n = 11,037
>18 yr
Average
individual born
-1969
Blood Pb (ICP-MS) (|jg/dL)f BP change
Quartiles (SBP, DBP)
Q1
Q2
Q3
Q4
<1.59
1.59-2.24
2.24-3.21
>3.21
Age at measurement
Range 18-79
Multiple linear regression
adjusted for sex, age, BMI,
regions, education, smoking
status, alcohol consumption,
family history of
hypertension, residence
area, rice consumption, red
meat consumptions,
vegetable consumptions,
FBG, TC, HDL-C, urinary
arsenic levels, and blood Cd
levels
BP Change (mmHg)b
SBP
Q2 vs. Q1 1.36 (0.25-2.47)
Q3 vs. Q1 1.38 (-0.25-3.00)
Q4 vs. Q1 4.72 (2.70-6.74)
DBP
Q2 vs.
Q1
1.06
(0.23-1.90)
Q3 vs.
Q1
1.94
(0.79-3.09)
Q4 vs.
Q1
4.42
(3.02-5.83)
Lopes et al.
(2017a)
Cambe, Brazil
2011
Cross-sectional
n = 948
Adults 40 yr
and older,
randomly
sampled from
census tracts in
the region
Average
individual born
-1956
Blood Pb (ICP-MS) (|jg/dL)
GM (95% CI):
1.97 (1.90-2.04)
10th percentile: 0.74
90th percentile: 6.03
Age at measurement:
Mean: 54.5 yr
BP (DBP, SBP)
Multiple linear regression
adjusted for age, sex, race,
income, education,
antihypertensive medication,
total cholesterol,
triglycerides, glycemia,
smoking, alcohol
consumption, and BMI
Change in BP (mmHg) (10th vs. 90th
percentile)
SBP no association (all Cis ranged
between 0 and 0)
DBP 0.005 (0.002, 0.008)
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Chung et al.
(2020)
Taiwan
n = 770 Blood Pb (ICP-MS) (pg/dL)
GM (IQR)
Distance from EAF
recruited 2010- arc furnace
2011 and 2015- (EAF)
2016
Community
residents living
<500 m: 2.41 (1.22-6.19)
near an electric v '
500-1000m: 2.26 (1.16-4.83)
1000-1500 m: 2.12 (1.05—
4.67)
1500-2000 m: 2.23 (0.98—
Average
Cross-sectional individual born 4.31)
-1953
>2000m: 2.03 (1.03-4.31)
BP (SBP, DBP) General linear models Change in BP (mmHg)b
adjusting for age, sex,
ethnicity, living near the main ggp- <| 43 (g 34 2 52)
road and smoking
DBP: 0.69 (0.01, 1.37)
Age at measurement:
Median 60
Wang et al.
(2020)
China
Cross-sectional
n = 816
Adults 40-75,
residing in area
for >15 yr, and
subsisting on
rice and
vegetables
grown in the
polluted (Cd
concentration
>0.2 mg/kg) or
unpolluted (Cd
concentration
<0.05 mg/kg)
area
Blood (ICP-MS) (|Jg/dL)f
Median (IQR)
Polluted 3.54 (2.42-4.89)
Unpolluted 2.61 (1.70-3.84)
Age at measurement:
mean (SD)
Polluted area
Hypertensive: 60.32 (8.08)
Normotensive: 55.61 (8.52)
Unpolluted area Hypertensive:
59.92 (9.19)
Normotensive: 56.86 (9.22)
BP (SBP, DBP)
Linear regression adjusted
forage, gender, smoking
status, and BMI
BP (mmHg) and Blood Pbb
See Figure 4-2 and Figure 4-3
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Zhang et al.
(2010)
Boston, MA
August 1991 and
December 2001
Cross-sectional
NAS
n = 619
Elderly men
(mostly white)
Average
individual born
-1933
Bone Pb (K-XRF) (pg/g)
Median (IQR)
Wild type HFE
Tibia: 8 (12-27)
Patella: 26 (17-37)
C282Y HFE
Tibia: 20 (14-27)
Patella: 25 (17-37)
H63D HFE
Tibia: 19 (14-26)
Patella: 27 (19-37)
Age at measurement
Mean: 67
BP (PP) Linear mixed-effects
regression models with
repeated measurements
adjusted for age, education,
alcohol intake, smoking, daily
intakes of calcium, sodium,
and potassium, total calories,
family history of
hypertension, diabetes,
height, heart rate, HDL, total
cholesterol, HDL ratio, and
waist circumference
PP (mmHg)
Tibia Pb
Wild Type HFE: 0.29 (-0.46, 1.05)
H63D HFE: 2.54 (0.12, 4.96)
C282Y HFE: 0.68 (-1.33, 2.70)
Any HFE variant: 2.23 (0.23, 4.23)
Patella Pb
Wild Type HFE: 0.14 (-0.33, 0.61)
H63D HFE: 1.53 (-0.005, 3.11)
C282Y HFE: 0.29 (-0.15, 0.73)
Any HFE variant: 1.49 (0.16, 2.82)
Weaver et al. n = 652 Blood Pb (GFAAS with
(2008) current and Zeeman correction) (pg/dL)
former Pb Mean (SD): 30.9 (16.7)
South Korea workers
Bone (Patella) Pb (K-XRF)
1999-2001 Average (pg/g)
individual born
-1957
Cross-sectional
Mean (SD): 75.1 (101.1)
Age at measurement:
Mean: 43.3, SD: 9.8
BP (SBP) Multivariable linear
regression adjusted for age,
sex, BMI, diabetes,
antihypertensive and
analgesic medication use, Pb
job duration, tobacco, and
alcohol use
SBP (mmHg)e
Blood Pb 0.1007 (0.02, 0.18)
Patella Pb 0.059 (-0.08, 0.20)
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Elmarsafawv et
al. (2006)
Boston, MA
Participants' first
visit after 1991
Cross-sectional
NAS
n = 471
Elderly men
(mostly white)
Average
individual born
-1924
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
Mean (SD): 6.6 (4.3)
Bone (K-XRF) (pg/g)
Mean (SD):
Tibia: 21.6 (12.0)
Patella: 31.7 (18.3)
Age at measurement
Mean: 67
BP (SBP) Multivariable linear
regression adjusted for age,
BMI, family history of
hypertension, smoking,
dietary sodium intake, and
cumulative alcohol ingestion
No control for potential
confounding SES factors
SBP (mmHg)
Tibia
Low Ca2+ group (<800 mg/d):
4.00 (1.05, 6.95)
High Ca2+ group (>800 mg/d):
1.90 (0.10, 3.70)
Yan et al. (2022) Haitian CVD Blood Pb (LeadCare II Blood BP (SBP, DBP)
Port-au-Prince,
Haiti
2019-2021
Cross-sectional
Cohort Study
n = 2,504
General
population >18
living in Port-
au-Prince
Average
individual born
-1980
Level Analyzer) (pg/dL)
(Pb measurement device had
high limit of detection
(3.3 pg/dL), and only 71% of
the population had
quantifiable blood Pb values)
GM: 4.73
Geometric SE: 1.62
Age at measurement
Median: 40 yr
Multivariable linear
regression models adjusted
forage, sex, BMI, smoking
status, alcohol use, physical
activity, income, and use of
antihypertensive medication
BP (mmHg)b
SBP
Q2 vs. Q1 0.62 (-1.46, 2.70)
Q3 vs. Q1 1.73 (-0.24, 3.70)
Q4 vs. Q1 2.42(0.36, 4.49)
>3.3 pg/dL vs. <3.3 pg/dL 1.65 (0.05,
3.24)
>5 pg/dL vs. <5 pg/dL 1.16 (-0.35, 2.68)
DBP
Q2 vs. Q1 0.19 (-1.26, 1.64)
Q3 vs. Q1 1.16 (-0.25, 2.57)
Q4 vs. Q1 1.96 (0.56, 3.37)
>3.3 pg/dL vs. <3.3 pg/dL 1.16 (0.04,
2.27)
>5 pg/dL vs. <5 pg/dL: 0.96 (-0.1, 2.02)
4-91
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Xu et al. (2021)
United States
2011-2013
Cross-sectional
GuLF Study
Cohort
n = 957
>21 yr
Average
individual born
in or before
-1991
Blood Pb (ICP-MS) (pg/dL)
Median: 0.09
75th percentile: 0.19
Maximum: 33.8
Age at measurement: >21 yr
BP (SBP, DBP)
Multiple linear regression
adjusted for age, sex, race,
educational attainment, and
household income
BP (mmHg)
SBP
Q2 vs. Q1 1.19 (-1.73, 4.11)
Q3 vs. Q1 0.54 (-2.48, 3.55)
Q4 vs. Q1 -0.96 (-4.13, 2.22)
DBP
Q2 vs. Q1 0.21 (-1.81, 2.24)
Q3 vs. Q1 0.26 (-1.84, 2.35)
Q4 vs. Q1 -0.01 (-2.21, 2.19)
Perlstein et al.
(2007)
Boston, MA
1991-1997
Cross-sectional
NAS
n = 593
Elderly men
(mostly white)
Average
individual born
-1927
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
Mean (SD)
Overall: 6.12 (4.03)
Q1
Q2
Q3
Q4
Q5
2.3 (0.8)
3.9 (0.3)
5.4 (0.5)
7.4 (0.6)
12.4 (4.4)
BP (PP) Multiple linear regression
adjusted for age, height,
race, heart rate, waist
circumference, diabetes,
family history of
hypertension, education level
achieved, smoking, alcohol
intake, fasting plasma
glucose, and ratio of total
cholesterol to HDL
cholesterol
PP (mmHg)b
Tibia Pb (above/below median): 4.2 (1.9,
6.50)
Tibia Pb (mean difference)
Q5 vs. Q1: 2.58 (-1.15, 6.48)
Blood Pb (mean difference)
Q5 vs. Q1: -1.49 (-4.93, 1.94)
Tibia (K-XRF) (pg/g)
Median: 19
Mean (SD)
Q1: 7.4 (3.2)
Q2: 14.1 (1.4)
Q3 18.9 (1.4)
Q4 24.9 (2.2)
Q5 40.9 (14)
4-92
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Ettinqer et al.
(2014)
Ghana, South
Africa,
Seychelles,
Jamaica, and the
United States
2010-2011
Cross-sectional
Modeling the
Epidemiologic
Transition
Study (METS)
n = 150 (30
randomly
selected from
each site)
Young adults
(25-45) of
African descent
Average
individual born
Blood Pb (ICP-MS) (pg/dL)
GM:
1.55 (95% CI: 1.30, 1.85);
Median
1.66 (95% CI: 1.34, 1.93)
75th: 2.6
Max: 31.82
Age at measurement (years):
Mean (SD):
Males: 34.7 (6)
Females: 35.2 (6.2)
BP (SBP, DBP)
Logistic regression adjusted
for age, sex, site location,
education, paid employment,
marital status, smoking,
alcohol use, fish intake
(percent body fat in models
not assessing models of
BMI)
OR above/below median (1.66 pg/dL)b
High SBP (>130 mmHg) 1.69 (0.55, 5.15)
High DBP (>85 mmHg) 2.20 (0.59, 8.16)
-1975
Guo et al. (2019) n = 145
China
2015
Cross-sectional
Males free of
cardiovascular
disease
(including
angina
pectoris, Ml,
and thrombus)
Average
individual born
-1976
Blood Pb (ICP-MS) (pg/dL)
mean (SD): 8.50 (3.77)
Median: 7.85
75th: 10.08
Max: 28.17
Age at measurement (years):
mean (SD): 39 (12)
BP (SBP, DBP)
Linear (log-transformed) and
logistic regression adjusting
for age
Linear regression (log-transformed)
(mmHg)b
SBP: 7.28 (-7.68, 22.24)
DBP: 4.34 (-7.12, 15.80)
OR above/below median (7.85 pg/dL)b
SBP >134 mmHg: 2.28 (1.01, 5.12)
DBP >84 mmHg: 1.55 (0.70, 3.40)
4-93
-------
Reference and
Study Design
Study
Population
Exposure Assessment Outcome
Confounders
Effect Estimates and 95% Clsa
Cohort Studies
Gambelunqhe et
Malmo Diet
Blood Pb (ICP-MS) (|jg/dL) BP (SBP, DBP)
Linear regression adjusted
BP differences (mmHg) (Q4 vs.
al. (2016)
and Cancer
Mean: 2.8
for sex, age, smoking,
Q1+Q2+Q3)be
Study (MDCS-
Max: 25.8
alcohol, waist circumference,
SBP: 1.8 (0.5, 3.1)
Malmo
CC)
education
Sweden
Quartile Means
SBP: Smokers 3.9 (1.6, 6.2)
n = 4,452
Q1: 1.5
SBP Never-smokers: 0.6 (-1.5, 2.7)
1991-1994, re-
Q2: 2.2
Males: 2.1 (0.3, 3.9)
examination
Aged 46-67
Q3: 2.8
Q4: 4.7
Females: 1.5 (-0.4, 3.4)
2007-2012
living in Malmo
Sweden
<57 yr: 2.4 (1.2, 3.6)
Cohort with
>57 yr: 1.3 (-0.6, 3.2)
cross-sectional
Average
Age at measurement (years):
component
individual born
-1935
Mean: 57
DBP: 1.4 (0.6, 2.2)
DBP Smokers: 1.6 (0.7, 2.5)
DBP Never-smokers: 1.1 (-0.05, 2.2)
Males: 1.7 (0.9, 2.5)
Females: 1.1 (0.2, 2.0)
<57 yr: 1.4 (0.7, 2.1)
>57 yr: 1.5 (0.75, 2.5)
4-94
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Bulka et al.
(2019)
Bangladesh
Participants were
randomized
between April
2006 and August
2009, BP
measurements
taken at baseline,
and every 2 yr for
a total of 6 yr
Cohort
Bangladesh
Vitamin E and
Selenium Trial
(BEST)
n = 255
Participants
from the trial
were from
randomized
sample of
those taking
part of the
placebo arm of
the BEST study
Blood Pb (whole) (ICP-MS)
(Hg/dL)
Median 8.5
Age at measurement
25-37 yr (88/255)
38-46 yr (82/255)
47-64 yr (85/255)
BP (SBP, DBP, Mixed-effects regression
PP) models adjusting for
manganese, selenium, age,
sex, site, smoking,
educational duration,
creatine-corrected urinary
arsenic, diabetes, BMI,
antihypertensive medication
use
Yearly change in BP (mmHg) (4th quartile
vs. 1st quartile)b
SBP: 1.16 (95% CI: 0.21, 2.11)
DBP: 0.53 (95% CI: -0.10, 1.16)
PP: 0.63 (95% CI: -0.08, 1.34)
Average
individual born
-1976, -1965,
-1952
Glenn et al.
(2006)
South Korea
1997-2001
Cohort
n = 575
Pb-exposed
workers
Average
individual born
-1956
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
Females
Visit 1
Visit 2
Visit 3
Males
Visit 1
Visit 2
Visit 3
20.3 (9.6)
20.8 (10.8)
19.8 (10.7)
35.0 (13.5)
36.5 (14.2)
35.4 (15.9)
BP (SBP) Multivariable models using
generalized estimating
equations were used in
longitudinal analyses
adjusted for visit number,
baseline age, baseline age
squared, baseline lifetime
alcohol consumption,
baseline BMI, sex, baseline
BP-lowering medication use,
alcohol consumption
Age at measurement:
Range 18-67 yr
SBP (mmHg)
Model 1 (short-term)
Blood Pb (longitudinal): 0.009 (0.002,
0.02)
Blood Pb (concurrent): 0.008 (-0.001,
0.02)
Model 4 (short and longer-term, controls
for tibia Pb)
Blood Pb (longitudinal): 0.009 (0.002,
0.02)
Blood Pb (concurrent): 0.01 (0.001, 0.019)
4-95
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Yu et al. (2020)
Belgium
Blood Pb
collected in
1985-2005,
arterial stiffness
measured a
median of 9.4 yr
later
Cohort
Cadmium in
Belgium study
n = 267
Average
individual born
-1958
Blood Pb (ETAAS) (pg/dL)
GM (IQR):
2.93 (1.8-4.7)
Age at measurement
Mean: 37 yr
Central and
Peripheral BP
Linear multivariable models
adjusting for sex, enrollment
characteristics (age, BMI,
smoking, drinking, serum
total to HDL-C ration, plasma
glucose, eGFR (estimated
from serum creatinine),
SES), the time interval
between measurement of
exposure biomarkers and
hemodynamic assessment,
and antihypertensive drug
treatment at enrollment and
follow-up
Per doubling of Pb concentration15
Peripheral
Systolic Pressure: 2.41 (-0.38, 5.20)
Diastolic Pressure: 0.50 (-1.07, 2.07)
PP: 1.91 (-0.32, 4.14)
Central
Systolic Pressure: 2.65 (-0.17, 5.46)
Diastolic Pressure: 0.42 (-1.18, 2.02)
PP: 2.23 (-0.03, 4.48)
Similar results when controlling for
baseline urinary Cd
AL = allostatic load; ALAD = 6-aminolevulinic acid dehydratase; BEST = Biomonitoring of Environmental Status and Trends; BMI = body mass index; BP = blood pressure; C282Y
HFE = mutant of the HFE wildtype; Ca2+ = calcium ion(s); Cd= cadmium; CI = confidence interval; CVD = cardiovascular disease; DBP = diastolic blood pressure; EAF = electric arc
furnace; eGFR = estimated glomerular filtration rate; ETAAS = electrothermal atomic absorption spectrometry; FBG = fasting blood glucose; GFAAS = graphite furnace atomic
absorption spectrometry; GM = geometric mean; GSE = geometric standard error; GuLF = Gulf Long-Term Follow-up; HDL-C = high-density lipoprotein cholesterol; H63D
HFE = mutant of the HFE wildtype; HFE = hemochromatosis gene; ICP-MS = inductively coupled plasma mass spectrometry; IQR = interquartile range; KNHANES = Korea National
Health and Nutrition Examination Survey; K-XRF = K-shell X-ray fluorescence; LF = low frequency; MAP = mean arterial pressure; MDCS-CC = cardiovascular cohort of the Malmo
Cancer and Diet Study; METS = Modeling the Epidemiologic Transition Study; Ml = myocardial infarction; mo = month(s); NAS = Normative Aging Study; NH = non-Hispanic;
NHANES = National Health and Nutrition Examination Survey; OR = odds ratio; Pb = lead; PIR = poverty-income ratio; PP = pulse pressure; Q = quartile; SBP = systolic blood
pressure; SD = standard deviation; SE = standard error; SES = socioeconomic status; TC = total cholesterol; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bUnable to be standardized.
°lncrement unclear.
dBlood Pb analysis method unclear, assumed based on data source.
Confidence intervals estimated based on reported p values.
'Original results reported in |jg/L.
4-96
-------
Table 4-4
Epidemiologic studies of Pb exposure and hypertension
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Teve et al. (2020)
United States
NHANES 1999-2016
Cross-sectional
NHANES
n = 30,467
20-79 yr
Average
individual born
-1965
Blood Pb ICP-MSC (pg/dL)
NH white men: 1.89
NH white women: 1.30
NH Black men: 2.20
NH Black women: 1.49
Hispanic men: 2.18
Hispanic women: 1.30
Other men: 1.93
Other women: 1.42
Hypertension
(SBP
>140 mmHg,
DBP >90, use of
antihypertensive
medication)
Race/ethnicity, age,
gender, education
level, BMI, and PIR
OR (95% CI):
1.002 (0.983, 1.021)
Age at measurement:
NH white men: 46.37
NH white women: 47.00
NH Black men: 43.09
NH Black women: 43.28
Hispanic men: 39.67
Hispanic women: 40.51
Other men: 42.92
Other women: 43.54
4-97
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Huanq (2022)
NHANES
Blood Pb (ICP-MS) (pg/dL)
Hypertension
Linear regression
Hypertension (OR)
n = 32,289
Mean (SD): 1.73 (1.71)
(SBP
models adjusted for
All 1.01 (0.99, 1.03)
United States
>20 yr
>130 mmHg,
DBP >80, use of
age, sex, race,
education, family
Age at measurement (years):
antihypertensive
income poverty
Women 1.03 (0.99, 1.07)
NHANES 1999-2018
Average
Mean (SD): 49.68 (18.04)
medication, or
ratio, BMI, alcohol
Men 1.01 (0.99, 1.03)
individual born
self-reported
use, and smoking
-1958
hypertension)
Cross-sectional
Mexican American 0.99 (0.96, 1.02)
Other Hispanic 1.01 (0.95, 1.06)
NH White 1.03 (1.00, 1.06)
NH Black 1.02 (0.98, 1.06)
Other race 1.04 (0.97, 1.11)
BMI >30 1.00 (0.97, 1.03)
BMI 25-30 1.00 (0.98, 1.03)
BMI <25 1.03 (1.00, 1.06)
4-98
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Tsoietal. (2021)
United States
NHANES 1999-2016
Cross-sectional
NHANES
n = 39,477
Adults >20
Average
individual born
-1960
Blood Pb ICP-MS (pg/dL)
Median: 1.30
Q1
Q2
Q3
Q4
<0.89
0.89-1.30
1.30-2.10
>2.10
Age at measurement
Mean (SE)
Hypertensive:
54.08 (0.23) yr
Non-hypertensive:
39.87 (0.19) yr
Hypertension
(SBP
>130 mmHg,
DBP >80, use of
antihypertensive
medication, or
self-reported
hypertension)
Multivariable logistic
regression adjusted
for age, sex
ethnicity, waist
circumference, PIR,
education, ever
cigarette smoking,
diabetes, and stage
3-5 chronic kidney
diseases
ORb
For every doubling of blood Pb
All 1.09 (1.04, 1.14)
Male 1.10 (1.05, 1.16)
Female 1.04 (0.97, 1.11)
Mex. American 0.98 (0.89, 1.08)
Other Hispanic 1.07 (0.93, 1.23)
NH White 1.12 (1.05, 1.19)
NH Black 1.06 (0.99, 1.15)
Other ethnicity 1.10 (0.95, 1.28)
Quartiles (All)
Q2 vs. Q1 1.15(1.04, 1.26)
Q3 vs. Q1 1.17(1.05, 1.31)
Q4 vs. Q1 1.21 (1.07, 1.36)
Miao etal. (2020)
United States
NHANES 1999-2016
Cross-sectional
NHANES
n = 30,762
>20 yr
Average
individual born
-1978, -1958, or
before -1947
Whole Blood Pb (GFAAS with
Zeeman correction) (1999-
2002) and ICP-MS (2003-
2016) (pg/dL)
mean (SE)
Male: 1.50 (0.02)
Female: 1.07 (0.01)
Age at measurement (years):
20-39 36.5%
40-59 38.9%
>60 24.6%
Uncontrolled
hypertension
(SBP
>130 mmHg or
DBP >80 mmHg
or
antihypertension
medication use)
and uncontrolled
hypertension
(SBP
>130 mmHg or
DBP >80 mmHg
regardless
antihypertension
medication use)
Logistic regression
adjusted for age,
sex, race/ethnicity,
ratio of family
income to poverty,
education, smoking
status, serum
cotinine, alcohol
intake, BMI, and
menopausal status
among females
OR (95% CI)
Uncontrolled hypertension vs. non-
hypertension:
Male: 1.037 (1.015, 1.060)
Female: 1.020 (0.970, 1.074)
Uncontrolled Hypertension vs.
Controlled Hypertension
Male: 1.157 (1.080, 1.239)
Female: 1.109 (1.020, 1.205)
Uncontrolled hypertension vs.
Controlled and Non-hypertension:
Male: 1.062 (1.036, 1.088)
Female: 1.056 (1.011, 1.102)
4-99
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Scinicariello et al. (2011)
United States
NHANES 1999-2006
Cross-sectional
NHANES
n = 16,222
Blood Pb
<10 |jg/dL
Average
individual born
-1959
Blood Pb (ICP-MS) (|jg/dL)
Percentile:
10th: <0.7
90th: 3.5-10
Age at measurement (years):
Mean (SE)
White males: 47.14 (0.37)
White females: 49.64 (0.36)
Black males: 42.86 (0.37)
Black females: 45.10 (0.42)
Mexican-Am males: 37.64
(0.48)
Mex-Am females: 40.67 (0.65)
Hypertension
prevalence (SBP
>140 mmHg,
DBP >90, use of
antihypertensive
medication)
Multivariable logistic
and linear
regression models.
Confounders: age,
BMI, self-reported
diabetes alcohol
ingestion, smoking
status, education,
serum creatinine,
serum total calcium,
sodium, hematocrit,
and blood Cd
Prevalence OR (90th vs. 10th
percentile)15
All: 1.26 (0.98, 1.61)
White males: 1.20 (0.74, 1.96)
White females: 1.07 (0.69, 1.66)
Black males: 2.69 (1.08, 6.72)
Black females: 1.04 (0.50, 2.16)
Mex-Am males: 1.03 (0.23, 4.59)
Mex-Am females: 0.67 (0.37, 1.20)
Hara et al. (2015)
United States
NHANES 2003-2010
Cross-sectional
NHANES
n = 12,725
>20 yr
Average
individual born
-1957
Blood Pb (ICP-MS) (|jg/dL)
GM (IQR):
Females
Black: 1.37 (0.88-2.10)
Hispanic: 1.21 (0.80-1.78)
White: 1.22 (0.80-1.86)
Males
Black: 1.86 (1.20-2.85)
Hispanic: 1.94 (1.25-2.83)
White: 1.73 (1.16-2.57)
Age at measurement:
mean (SD)
Black females: 48.31 (6.8)
Hispanic females: 48.1 (16.8)
White females: 53.0 (8.4)
Black males: 47.7 (16.9)
Hispanic males: 46.1 (6.8)
White males: 53.1 (18.6)
Hypertension
(SBP
>140 mmHg or
DBP >90 mmHg
or the use of
antihypertensive
medication)
Logistic models
adjusted for
ethnicity, sex, age,
BMI, heart rate,
hematocrit, serum
total calcium y-
glutamyltransferase,
cotinine, dietary
sodium to
potassium intake
ratio, college
education,
antihypertensive
drug treatment
OR (95% CI) for doubling blood Pbb
All: 0.95 (0.90, 1.01)
Females: 0.95 (0.87, 1.04)
Black: 0.82 (0.67, 0.99)
Hispanic: 0.86 (0.72, 1.04)
White: 1.06 (0.94, 1.21)
Males: 0.95 (0.87, 1.02)
Black: 1.00 (0.84, 1.20)
Hispanic: 0.84 (0.71, 0.99)
White: 0.99 (0.89, 1.10)
4-100
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Lee etal. (2016b)
Korea
KNHANES
2007-2013
Cross-sectional
KNHANES
n = 8,493
>20 yr
Average
individual born
-1981, -1961, or
before -1950
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
Quartiles
Q1
Q2
Q3
Q4
0.206-1.539
1.540-2.056
2.057-2.716
2.717-24.532
Age at measurement:
20-39 (58.8%)
40-59 (34.9%)
>60 (6.3%)
Prehypertension
prevalence (DBP
80-89 mmHg or
SBP 120-
139 mmHg and
the absence of
any current
treatment or
diagnosis of
hypertension)
Logistic regression
adjusted for age,
sex, education,
occupation, income
residence, smoking
alcohol
consumption,
exercise level,
serum creatinine
clearance, chronic
disease, and
antihypertensive
medication
ORb (95% CI):
Q2 vs. Q1 1.24
(1
.04,
1.48)
Q3 vs. Q1 1.27
(1
.06,
1.52)
Q4 vs. Q1 1.30
(1
.07,
1.60)
Lee etal. (2016a)
KNHANES
2008-2013
Cross-sectional
KNHANES
n = 11,797
>19 yr
Average
individual born in
or before -1991
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
GM (95% CI)
Male: 2.396 (2.362, 2.430)
Female: 1.919 (1.889, 1.949)
Hypertension
(SBP
>140 mmHg or
DBP >90 mmHg)
and
Prehypertension
(SBP
>120 mmHg or
DBP >80 mmHg)
Logistic models
adjusted for sex,
age, residence
area, education
level, smoking,
drinking status,
BMI, physical
activity, serum
creatinine, and
hemoglobin
OR for doubling blood Pbb
Hypertension
All: 1.09 (0.98, 1.22)
Male: 0.92 (0.80, 1.07)
Female: 1.29 (1.10, 1.51)
Prehypertension
All: 1.09 (0.99, 1.21)
Male: 0.98 (0.85, 1.12)
Female: 1.21 (1.06, 1.38)
Choi etal. (2018)
South Korea
KNHANES 2013
Cross-sectional
KNHANES
n = 1,350
19-64 yr
Average
individual born
-1989, -1978,
-1968, -1958,
-1951
Blood Pb (GFAAS with
Zeeman correction) (ug/dL)
Mean 2.01
SE 0.025
Age at measurement:
19-29 (23.1%)
30-39 (22.7%)
40-49 (21.9%)
50-59 (22.1%)
60-64 (10.1%)
Hypertension Logistic regression ORd (95% CI)
prevalence (SBP
>140 mmHg, or
DBP >90 mmHg,
or use of
antihypertensive
medication)
adjusting for age,
sex, smoking, and
BMI
Curry intake:
1.108 (0.827, 1.484)
Non-curry intake:
1.399 (1.054, 1.857)
4-101
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Qu et al. (2022)
China
2017-2018
Cross-sectional
China National
Human
Biomonitoring
Study
n = 11,037
18-79 yr
Average
individual born
-1969
Blood Pb (ICP-MS) (|jg/dL)e
Quartiles
Q1
Q2
Q3
Q4
<1.59
1.59-2.24
2.24-3.21
>3.21
Prehypertension
(SBP 120-139,
DBP 80-89),
hypertension
(Chinese
guideline: SBP
>140, or DBP
>90), elevated
BP (2017
ACC/AHA: SBP
120-129, DBP
<80), stage 1
hypertension
(2017 ACC/AHA:
SBP 130-139,
DBP 80-89)
Logistic regression
adjusted for sex,
age, BMI, regions,
education, smoking
status, alcohol
consumption, family
history of
hypertension,
residence area, rice
consumption, red
meat consumptions,
vegetable
consumptions,
FBG, TC, HDL-C,
urinary arsenic
levels, and blood
Cd levels
ORb (95% CI)
Prehypertension
Q2 vs. Q1
Q3 vs. Q1
Q4 vs. Q1
1.24 (1.04-1.47)
1.27 (1.02-1.59)
1.56 (1.22-1.99)
Hypertension
Q2 vs. Q1
Q3 vs. Q1
Q4 vs. Q1
1.23 (0.96-1.56)
1.49 (1.12-1.96)
2.33 (1.67-3.24)
Elevated BP
Q2 vs. Q1
Q3 vs. Q1
Q4 vs. Q1
1.00 (0.77-1.31)
1.10 (0.83-1.47)
1.18 (0.88-1.57)
Stage 1 Hypertension
Q2 vs. Q1: 1.35 (1.10-1.65)
Q3 vs. Q1: 1.35 (1.04-1.77)
Q4 vs. Q1: 1.75 (1.31-2.33)
Bushnik et al. (2014)
Canada
2007-2011
Cross-sectional
Canadian Health Blood Pb (ICP-MS) (|jg/dL)
Measures Survey Mean: 1.64 (1.58-1.71)
n = 4,550
Nonpregnant
individuals aged
40-79
Average
individual born
-1954
Age at measurement:
Mean: 55.4
Hypertension:
SBP >140, or
DBP >90, or use
of
antihypertensive
medication, or
health care
provider
diagnosis of
hypertension
Logistic regression
adjusted for age,
sex, education,
smoking, alcohol,
physical activity,
BMI, non-HDL
cholesterol,
diabetes, chronic
kidney disease,
family history of
high BP,
antihypertension
medication use
OR
Age 40-79: 0.02 (0.00, 0.43)
Age 40-54: 0.01 (0.00, 0.20)
Age 55-79: 0.01 (0.00, 1.00)
Male: 0.00 (0.00, 0.91)
Female: 0.02 (0.00, 0.60)
*Results correspond to linear model.
Concentration response function for
splines not shown. Authors indicated
no relationship between hypertension
and BLL
4-102
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Wang et al. (2020)
China
Cross-sectional
n = 816
Adults 40-75,
residing in area
for >15 yr, and
subsisting on rice
and vegetables
grown in the
polluted (Cd
concentration
>0.2 mg/kg) or
unpolluted (Cd
concentration
<0.05 mg/kg)
area
Blood Pb (ICP-MS) (|jg/dL)
Median (IQR)
Polluted: 3.54 (2.4-4.89)
Unpolluted: 2.61 (1.70-3.84)
Age at measurement:
Mean (SD)
Polluted area
Hypertensive: 60.32 (8.08)
Normotensive 55.61 (8.52)
Unpolluted area Hypertensive:
59.92 (9.19)
Normotensive 56.86 (9.22)
Hypertension
(SBP
>140 mmHg,
DBP >90 mmHg,
self-reported
physician
diagnosis, or
current use of
antihypertensive
medication)
Logistic regression
adjusted for age,
gender, smoking
status, and BMI
ORb (95% CI)
See Figure 4-2 and Figure 4-3
Lopes et al. (2017a)
Cambe, Brazil
2011
Cross-sectional
n = 948
adults 40 yr and
older, randomly
sampled from
census tracts in
the region
Average
individual born
-1956
Blood Pb (ICP-MS) (pg/dL)
GM:
1.97 (95% Cl:1.90-2.04)
Age at measurement:
Mean: 54.5 yr
Hypertension
(SBP
>140 mmHg,
DBP >90 mmHg
or current
antihypertensive
medication)
Logistic regression
adjusted for age,
sex, race, income,
education,
antihypertensive
medication, total
cholesterol,
triglycerides,
glycemia, smoking,
alcohol
consumption, and
BMI
OR 1.079 (1.026, 1.136)
4-103
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Xu et al. (2021)
United States
2011-2013
Cross-sectional
GuLF Study
Cohort
n = 957
Adults >21
Average
individual born in
or before -1991
Blood Pb (ICP-MS) (pg/dL)
Median: 0.09
75th percentile: 0.19
Maximum: 33.8
Age at measurement: >21 yr
Hypertension
(SBP
>140 mmHg,
DBP >90 mmHg
or current
antihypertensive
medication)
Multivariable
Poisson regression
adjusted for age,
sex, race,
educational
attainment, and
household income
Prevalence Ratio (PR)
Q2 vs. Q1
Q3 vs. Q1
Q4 vs. Q1
BMI >30
Q2 vs. Q1
Q3 vs. Q1
Q4 vs. Q1
0.96 (0.73,1.25)
0.91 (0.71, 1.17)
0.86 (0.66, 1.12)
1.05 (0.78, 1.42)
1.09 (0.82, 1.45)
1.14 (0.84, 1.55)
BMI <30
Q2 vs. Q1
Q3 vs. Q1
Q4 vs. Q1
0.89 (0.52, 1.52)
0.81 (0.50, 1.31)
0.82 (0.50, 1.32)
Weaver et al. (2008)
South Korea
1999-2001
Cross-sectional
n = 652
current and
former Pb
workers
Average
individual born
-1957
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
Mean (SD): 31.9 (14.8)
Patella Pb (K-XRF) (pg/g)
Mean (SD): 75.1 (101.1)
Age at measurement:
Mean: 43.3, SD: 9.8
Hypertension
(SBP
>140 mmHg,
DBP >90 mmHg;
and/or use of
antihypertensive
medications; or
physician
diagnosis)
Logistic regression
models adjusted for
age, sex, BMI,
diabetes,
antihypertensive
and analgesic
medication use, Pb
job duration, work
status, tobacco, and
alcohol use
Quantitative results not reported.
None ofthe examined Pb exposure
metrics (blood, patella, and
logarithmic (In)-transformed patella)
were significantly associated with
hypertension
4-104
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Elmarsafawv et al. (2006)
Boston, MA
1991
Cross-sectional
NAS
n = 471
Elderly men
(mostly white)
Average
individual born
-1924
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
Mean (SD): 6.6 (4.3)
Bone (K-XRF) (pg/g)
Mean (SD)
Tibia: 21.6 (12.0)
Patella: 31.7 (18.3)
Age at measurement
Mean: 67
Hypertension
(SBP
>160 mmHg,
DBP >95 mmHg;
and/or physician
diagnosis with
current use of
antihypertensive
medications)
Logistic regression
models adjusted for
age, BMI, family
history of
hypertension,
history of smoking,
dietary sodium
intake, and
cumulative alcohol
ingestion
OR (95% CI)
Low Ca2+ group (<800 mg/d):
Blood: 1.02 (1.00, 1.03)
Tibia: 1.02 (1.00, 1.03)
Patella: 1.01 (1.00, 1.01)
High Ca2+ group (>800 mg/d):
Blood Pb: 1.01 (0.99, 1.02)
Tibia Pb: 1.02 (1.00, 1.05)
Patella Pb: 1.01 (1.00, 1.02)
Gambelunahe et al. (2016) Malmo Diet and Blood Pb (ICP-MS) (pg/L)
Malmo, Sweden
1991-1994, re-examination
2007-2012
Cohort
Cancer Study
(MDCS-CC)
n = 4,452
aged 46-67
living in Malmo
Sweden
Average
individual born
-1935
All: 2.8
Max: 25.8
Age at measurement:
Mean: 57
Range: 46-67
Hypertension
(SBP
>140 mmHg, or
DBP >90 mmHg,
or
antihypertensive
medication use)
Logistic regression OR (Q4 vs. Q1+Q2+Q3)b
adjusted for sex,
age, smoking,
alcohol, waist
circumference,
education
All: 1.3 (1.1, 1.5)
Smokers: 1.5 (1.2, 1.8)
Never-smokers: 0.96 (0.7, 1.3)
<57 yr: 1.5(1.2, 1.9)
>5 yr: 1.3 (0.9, 1.4)
Male: 1.20 (0.60, 1.5)
Female1.4 (1.1, 1.7)
At Follow-up
Antihypertensive medication use: 1.0
(0.8, 1.2)
High BP: 1.0 (0.7, 1.3)
4-105
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Zheutlin et al. (2018)
Boston, MA
1986-2013
Cohort
NAS
n = 475
Male volunteers
aged 21 to 80 yr
(bone Pb
measurement
started in 1991,
resistant
hypertension
assessed starting
at visit prior to
first bone
measurement)
Average
individual born
-1931
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
Median (IQR): 5.0 (3.4-8.0)
Bone (K-XRF) (ug/g)
Median (IQR)
Tibia: 20.0 (13.0-28.5) Patella:
27.0 (18.0-40.0)
Age at measurement:
67.9 (63.2-72.6)
Incident resistant
hypertension
(inadequate
control) (SBP
>140 mmHg,
DBP >90 mmHg)
while taking >3
antihypertensive
medications, or
adequate control
(SBP
<140 mmHg,
DBP <90 mmHg)
while taking >4
antihypertensive
medications
Poisson regression
with robust error
variation adjusting
for age,
race/ethnicity,
education
attainment, income
level, BMI, family
history of
hypertension, and
cigarette smoking
RR
Tibia: 1.12 (1.01, 1.25)
Patella: 1.04 (0.96
Blood: 1.02 (0.97,
1.13)
.08)
ACC = American College of Cardiologists; AHA = American Heart Association; BLL = blood lead level; BMI = body mass index; BP = blood pressure; Ca2+ = calcium ion;
Cd = cadmium; CI = confidence interval; DBP = diastolic blood pressure; FBG = fasting blood glucose; GFAAS = graphite furnace atomic absorption spectrometry; GM = geometric
mean; GuLF = Gulf Long-Term Follow-up; HDL-C = high-density lipoprotein cholesterol; ICP-MS = inductively coupled plasma mass spectrometry; IQR = interquartile range;
KNHANES = Korea National Health and Nutrition Examination Survey; LF = low frequency; MDCS-CC = cardiovascular cohort of the Malmo Cancer and Diet Study; Mex-
Am = Mexican-American; NAS = Normative Aging Study; NH = non-Hispanic; NHANES = National Health and Nutrition Examination Survey; OR = odds ratio; PIR = poverty-income
ratio; Pb = lead; Q = quartile; PR = prevalence ratio; RR = relative risk; SBP = systolic blood pressure; SD = standard deviation; SE = standard error; TC = total cholesterol; K-
XRF = K-shell X-ray fluorescence; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bUnable to be standardized.
°Blood Pb analysis method unclear, assumed based on data source.
Confidence intervals estimated based on reported p values.
eOriginal results reported in |jg/L.
4-106
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Table 4-5
Epidemiologic studies of Pb exposure and blood pressure and hypertension among children
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Zhang et al. (2012)
Mexico City, Mexico
1994-2003
Follow-up 2008-2010
Cohort
Early Life Exposures Cord Blood (AAS) (pg/dL) BP (SBP, DBP) in
in Mexico to
Environmental
Toxicants project
n = 457 mother-child
pairs
Average individual
born -1973
Mean (SD): 5.51 (3.45)
Bone (K-XRF) (pg/g)
Median (IQR)
Tibia: 9.3 (3.3-16.1)
Patella: 11.6 (4.5-19.9)
Age at measurement
Mean (SD): 25.6 (5.4)
children
Age at outcome
Mean (SD): 10.7
(2.4)
Multiple
regression models
and generalized
estimating
equations (log
linear for cord
blood, linear for
concurrent blood
and maternal
bone) adjusted for
maternal
education, birth
weight, BMI, sex,
and child
concurrent age
Difference in BP (mmHg)
Cord Blood
SBP All: 0.23 (-0.14, 0.60)
DBP All: 0.23 (-0.03 0.49)
Tibia
SBP All: 0.74 (-0.10, 1.58)
SBP Female: 1.62 (0.54, 2.71)
SBP Male: -0.26 (-1.52, 1.00)
DBP All: 0.35 (-0.36, 1.07)
DBP Female: 1.24 (0.23, 2.25)
DBP Male: -0.64 (-1.57, 0.30)
Patella
SBP All: 0.28 (-0.45, 1.00)
DBP All: 0.14 (-0.59, 0.88)
Gump et al. (2005)
Oswego, NY (born at a
single hospital in New York
from 1991-94)
Cohort
Oswego Children's
Study
n = 122 children aged
9.5 yr
Average individual
born -1990
Cord blood (ETAAS)
(pg/dL)
Mean (SD): 2.97 (1.75)
Child blood (ETAAS and
ASV) (pg/dL)
Mean (SD): 4.62 (2.51)
BP (SBP, DBP) and
TPR in children
Age at outcome 9.5
Linear regression
models examined
the adjusted for
HOME score, SES,
birth weight, child
BMI, and child sex
Baseline BP (mmHg) per 1 pg/dL
increase in cord blood Pbb
SBP: 12.16 (2.44, 21.88)
DBP: 8.45 (-0.45, 17.35)
TPR, dyne-s/cm5
no association, results NR
Age of child blood Pb
measurement
Mean: 2.6
Relationship of blood Pb with
change in z-score for outcome
(post and prestress) per 1 pg/dL
increase in childhood blood Pb
SBP: -0.009 (-0.074, 0.055)
DBP: 0.069 (-0.001, 0.138)
TPR, dyne-s/cm5
0.088 (0.024, 0.152)
4-107
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Gump et al. (2007)
Oswego, NY (born at a
single hospital in New York
from 1991-94)
Cohort
Oswego Children's
Study
n = 122 children aged
9.5 yr
Average individual
born -1990
Cord blood (ETAAS)
(pg/dL)
Mean (SD): 2.97 (1.75)
Child blood (ETAAS and
ASV) (pg/dL)
Mean (SD): 4.62 (2.51)
Age of child blood Pb
measurement (years):
Mean: 2.6
BP (SBP) and TPR
in children
Age at outcome 9.5
Linear regression
models adjusting for
the same covariates
as in Gump et al.
(2005). Separate
models testing
whether Pb is a
mediator of SES
associations (Sobel
test) and whether Pb
moderates SES
associations (Pb-
SES interaction)
Blood Pb was a mediator of the
SES-TPR relationship
SES alone: -0.62 dyne-s/cm5
(p < 0.05)
SES with Blood Pb: -0.40 dyne-
s/cm5 (p > 0.10), change in R2
attributable to SES: -55.3%
Blood Pb was a potential
moderator of the SES-TPR
relationship.
Blood Pb x SES interaction:
p = 0.07
Blood Pb was a moderator of
SES-SBP relationship
Pb x SES interaction: p = 0.007
Kupsco et al. (2019)
Mexico City, Mexico
2007-2011
Cohort
Research in Obesity, Maternal Blood (ICP-MS) BP (SBP, DBP)
Growth, Environment
and Social Stressors
(PROGRESS) birth
study
n = 548
Mother/child pairs
Maternal blood
tested for metals in
second trimester,
children assessed at
age 4-6
(pg/dL)
Mean: 3.7, SD: 2.7
Range 0.75-18
Max: 18
Age at measurement
(years):
28 (5.6)
Age at Outcome:
Mean (SD): 4.8
(0.55)
Range: 4-6.8
Linear regression
adjusted for birth
weight, gestational
age, prepregnancy
BMI, education,
socioeconomic
status, parity,
environmental
tobacco smoke
BP (mmHg) per 1 In unit increase
in maternal blood Pbc
SBP: -0.05 (-0.09, 0.07)
DBP: 0 (-0.23, 0.54)
Average individual
born -1981
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Skroder et al. (2016)
Bangladesh
(2002- 2004)
Cohort
Maternal and Infant
Nutrition
Interventions, Matlab
n = 1,511
(gestational week
[GW] 14); 713 (GW
30)
Mother-child pairs
Eyr-Pb measured at
GW14 and GW30,
children assessed at
age 4.5
Maternal Blood (Eyr-Pb)
(ICP-MS) (pg/kg)
GW: 14
Median: 73
95th: 172
GW: 30
Media: 86
95th : 506
Age at measurement:
Mean (SD): 26 (6)
BP (SBP, DBP)
Age at Outcome
Mean: 4.5 yr
Linear regression
adjusted for sex,
birth weight, season
of birth, age at
outcome
measurements,
height for age z-
score, maternal BMI
at GW8, parity, SES,
and supplementation
group
BP (mmHg) per pg/kg Eyr-Pbc
GW14:
SBP: 0.042 (-0.058, 0.14)
DBP: -0.0058 (-0.090, 0.077)
GW30:
SBP: 0.042 (
DBP: 0.072 I
-0.090, 0.17)
-0.039, 0.18)
Average individual
born -1977
Gump et al. (2011)
Oswego, NY
Cross-sectional
n = 140 children
ages 9-11 yr
Blood (ICP-MS) (pg/dL)
GM: 1.01
BP (SBP), TPR
Q1
Q2
Q3
Q4
0.14-0.68
0.69-0.93
0.94-1.20
1.21-3.76
Linear regression
adjusted for sex,
SES, BMI, and age
Change in SBP (mmHg) across
quartiles in response to acute
stresscd
Q1: 5.30, Q2: 7.33, Q3: 7.07, Q4:
7.23, p for trend = 0.31
Change in TPR (%) across
quartiles in response to acute
stresscd
Q1: 2.91, Q2: 8.18, Q3: 9.55, Q4:
9.51, p for trend = 0.03
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Factor-Litvak et al. (1996)
Kosovo (when part of
Yugoslavia)
January 1st, 1984-July 31st,
1986
Cross-sectional
Yugoslavia
Prospective Study
n = 281 from two
towns (Kovoska
Mitrovica [Exposed
town] and Pristina
[Unexposed town
25 miles south])
Average individual
born -1980
Blood (GFAAS) (pg/dL) BP (SBP, DBP)
Exposed Town
Mean (SD): 37.3 (12.0)
Range: 9.5- 76.4
Unexposed Town
Mean (SD): 8.7 (2.8)
Range: 4.1-20.2
Linear regression
adjusted for ethnic
group, birth order
(and height, BMI,
and sex for SBP and
waist circumference
for DBP)
BP (mmHg) per 1 pg/dL blood Pb
SBP: 0.054 (-0.024, 0.13)
DBP: 0.042 (-0.010, 0.090)
Lu et al. (2018)
Guiyu (e-waste exposed),
Haojiang (reference)
China
2016
Cross-sectional
n = 590
children (aged 3-7)
residing in either
Guiyu or Haojiang
China
Average individual
born -2011
Blood (GFAAS) (pg/dL)
Median
Exposed: 7.14
Unexposed: 3.91
Age at measurement
Mean (SD)
Exposed: 4.52 (0.86)
Unexposed 4.40 (1.04)
BP (SBP, DBP)
Age at Outcome
Mean (SD)
Exposed: 4.52
(0.86)
Unexposed: 4.4
(1.04)
Linear regression
adjusted for outdoor
activities, family
member smoking,
parent education and
diet (including
cooking oil, picky
eating, sweetmeat
consumption, salted
products, vegetable
and fruit
consumption, dairy
products, bean
products, marine
products), age, sex,
BMI, and family
history of diseases
(hypertension,
diabetes, obesity)
Ln-transformed Blood Pb and BP
(mmHgf
SBP: -0.30 (-2.6, 2.00)
DBP: -2.11 (-4.38, 0.17)
4-110
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Ahn etal. (2018)
Korea
KNHANES
2010-2016
Cross-sectional
KNHANES
n = 1,776
Adolescents (10-
18 yr)
Average individual
born -1999
Blood (GFAAS with
Zeeman background
correction (pg/dL)
GM (95% CI): 1.19 (1.17-
1.22)
Age at measurement:
Range: 10-18 yr
BP (SBP, DBP)
Prehypertension
(SBP 120-
140 mmHg, DBP
80-90 mmHg)
Linear and logistic
regression adjusted
for sex, age,
residence area,
smoking status,
drinking status, BMI,
year of
measurement,
physical activities,
hemoglobin, and
serum creatinine
Mean difference (mmHg) for
doubling blood Pbc
DBP: -0.680 (-1.561, 0.221)
SBP: -0.0999 (1.098, 0.898)
Prehypertension (OR [95% CI]) for
doubling blood Pbc
0.906 (0.629, 1.305)
Xu et al. (2017)
United States
NHANES 1999-2012
Cross-sectional
NHANES
n = 11,662
Adolescents 12-19
participating in
NHANES
Average individual
born -1990
Blood (ICP-MS)C (|jg/dL)
Mean (SD) 1.17 (1.20)
Q1: <0.6
Q2: 0.6-0.9
Q3: 0.0-1.34
Q4: >1.34
Age at measurement
Range: 12-19 yr
BP (SBP, DBP)
Linear models
adjusted for age,
sex, PIR, waist
circumference,
serum cotinine,
physical activity and
NHANES cycle
BP (mmHg) (Q4 vs. Q1)c
SBP: 0.001 (-0.001, 0.004)
DBP: 0.001 (-0.006, 0.008)
4-111
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Yao et al. (2020)
United States
2007-2016
Cross-sectional
NHANES
n = 7,076
Children and
adolescents
Average individual
born -1999
Blood (ICP-MS) (ug/dL)
Mean: 0.67
Q1
Q2
Q3
Q4
<0.46
0.46-0.65
0.65-0.96
>0.96
Age at measurement
Mean (SD): 11.99 (2.88)
BP (SBP, DBP)
High BP (self (or
parent) reported
hypertension
diagnosis or
antihypertension
medication use for
those >16, or SBP
>120 mmHg or
DBP >80)
Linear or logistic
regression adjusted
for age, sex,
race/ethnicity, BMI,
cycle, serum cotinine
levels, hematocrit,
annual family
income, and intake
of calcium, sodium,
and potassium
BP (Regression coefficient) for
log-transformed blood Pbbc
SBP
All: -0.48 (-1.07, 0.11)
Male: -0.55 (-1.35, 0.25)
Female: -0.53 (-1.41, 0.35)
Mexican American: -0.10 (-1.06,
0.86)
Other Hispanic: -1.77 (-3.46,
-0.08)
White: -0.27 (-1.21, 0.67)
Black: 0.17 (-0.15, 1.65)
DBP
All: 0.75 (-1.01, 1.49)
Male:1.16 (-0.13, 2.45)
Female: 0.24 (-1.01, 1.49)
Mexican American: 1.12 (-0.49,
2.73)
Other Hispanic: -0.86 (-2.53,
0.81)
White: 1.99 (0.58, 3.40)
Black: -2.30 (-4.38, -0.22)
High BP (OR) Q4 vs. Q1
All: 0.89 (0.62-1.27)
"Change in BP associated with a
twofold increase in blood Pb was
calculated by dividing the
regression coefficient by log2€
4-112
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Desai et al. (2021)
NHANES
2009-2016
Cross-sectional
NHANES
n = 1,642
participants aged 8-
17 yr
Average individual
born -2000
Blood Pb (ICP-MS) (pg/dL)
Median: 0.57
95th percentile: 1.6
Age at measurement
Median: 152 mo (-12.7 yr)
BP (SBP, DBP, PP)
Multivariate linear
regression models
adjusted for age,
sex, race, BMI, total
energy intake,
NHANES cycle,
education of
household head, and
income to poverty
ratio
BP (mmHg)
SBP: -0.351 (-1.391, 0.689)
DBP: -0.078 (-1.365, 1.209)
PP: -0.273 (-1.781, 1.235)
Zhang et al. (2021)
Boston, MA (United States)
Baseline 2002-2013, follow-
up through 2018
Cohort
Boston Birth Cohort
n = 1,194 mother-
child pairs
Average individual
born -1980
Maternal red blood Pb
(measured 24-72 hr
postdelivery, ICP-MS)
(pg/dL)
Median: 2.42
75th percentile: 3.68
Max: 24.8
Age at measurement
(age at delivery)
Mean (SD): 27.7 (6.5)
BP (SBP) percentile
(based on child
age, sex, and
height according to
the 2017 American
Academy of
Pediatric Clinical
Practice Guideline)
Age at outcome
Median (IQR):
8.4 (6.2-10.6)
Multivariate linear
regression models
adjusting for
maternal age,
race/ethnicity,
educational level,
prepregnancy BMI,
and smoking history
Difference in child SBP percentile0
Quartiles
Q2 vs. Q1: 0.92 (-3:13, 4.97)
Q3 vs. Q1: 1.39 (-2:85, 5.64)
Q4 vs. Q1: -0.62 (-4.97, 3.74)
Per 1 pg/dL increase in blood Pb
0.142 (-0.673, 0.958)
AAS = atomic absorption spectrometry; ASV = anodic stripping voltammetry; BMI = body mass index; BP = blood pressure; CI = confidence interval; DBP = diastolic blood pressure;
ETAAS = Electrothermal Atomic Absorption Spectrometry with Zeeman background correction; Eyr = erythrocyte; GFAAS = graphite furnace atomic absorption spectrometry;
GM = geometric mean; GSD = geometric standard deviation; GW = gestational week; HOME = Health Outcomes and Measures of the Environment; ICP-MS = inductively coupled
plasma mass spectrometry; IQR = interquartile range; KNHANES = Korea National Health and Nutrition Examination Survey; mo = month(s); Mex-Am = Mexican American;
NHANES = National Health and Nutrition Examination Survey; NR = not reported; OR = odds ratio; PIR = poverty-income ratio; PP = pulse pressure; PROGRESS = Programming
Research in Obesity, Growth, Environment and Social Stressors; Q = quartile; SBP = systolic blood pressure; SD = standard deviation; SES = socioeconomic status; TPR = total
peripheral resistance; Pb = lead; yr = year(s).
aEffect estimates are standardized to a 1 pg/dL increase in blood Pb or a 10 pg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th~90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bConfidence intervals estimated based on reported standard errors.
°Unable to be standardized.
Confidence intervals not provided and unable to calculate based on given information.
4-113
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Table 4-6 Animal toxicological studies of Pb exposure and blood pressure/hypertension
Study (Stock/Strain), Exposure Details Reported (pg/dL) Endpoints Examined
n, Sex Exposure (Concentration, Duration) ^ ^
Fioresi et al. Rat (Wistar) Age 2 mo to 3 mo 100 ppm Pb acetate in drinking <0.5 |jg/dL for control SBP, measured weekly from 0 to
(2014) Control (tap water for 30 d 4wk
water), M, 13.6 ± 1.07 |jg/dL for MAP, DBP, ACE activity measured
n = 9-12 100 ppm group post 30-d exposure
100 ppm
group, M,
n = 9-12
Gaspar and
Cordellini (2014)
Rat (Wistar)
Control (tap
water), M,
n = 15
8500 ppm
group, M,
n =20
In utero to PND 22
Pregnancy day 0: females
divided into tap water and
500 ppm Pb acetate in drinking
water groups. Exposure lasted
through pregnancy. At birth,
pups were exposed to Pb (or
control) through nursing. Pups
were weaned at 22 d
<5 |jg/dL at all times for
tap water
19.98 ±6.31 - PND 52
13.15 ± 0.97 - PND 70
11.17 ± 2.11 -PND 100
SBP weekly measurements starting
at PND 23 to PND 100
Nunes et al. (2015) Rat (Wistar) 2 mo old rats 100 ppm Pb acetate in drinking NR for control SBP measured post 30 d exposure
Control exposed for 30 d water for 30 d for 28 d
(Distilled 8.4 |jg/dL for 100 ppm
water),
M, n = 5
100 ppm Pb
acetate group,
M, n = 5
4-114
-------
Study
Species
(Stock/Strain),
n, Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (pg/dL)
Endpoints Examined
Silva etal. (2015)
Rat (Wistar)
Control
(distilled water)
M, n = 6
3 mo old rats 100 ppm Pb acetate in drinking 12.3 ± 2 [jg/dL
exposed for 15 d water for 15 d
SBP measured weekly
100 ppm Pb
acetate group
M, n = 6
Wildemann et al.
(2015)
Rat (Wistar)
Control (tap
water), M,
n =6
357 or
1607 |jg/kg
BW/d, M, n = 5
per group
Unknown start age
for 4 wk
Tap water with 0.2% nitic acid,
or 357 or 1607 |jg/kg BW/d Pb
acetate in drinking water for
4 wk
1.4 ± 1.2 |jg/L for tap water/
0.2% nitic acid
(0.14 ± 0.12 pg/dL)
17 ± 7 pg/L for 357 pg/kg
BW/d Pb acetate
(1.7 ± 0.7 pg/dL)
86 ±29 pg/L for 1607 pg/kg
BW/d Pb acetate
(8.6 ±2.9 pg/dL)
DBP, PP, SBP all measured post 4-
wk exposure
Xu et al. (2015)
Rat (Sprague
6-7-wk-old rats
Distilled water for 40 d or 1%
Day 12:
DBP, SBP, measured intermittently
Dawley)
exposed for 12 or
Pb acetate in drinking water
193.3 pg/L (19.33 pg/dL)
from 0-40 d
Control 1, M/F,
40 d
for 12 or 40 d
n =6
Day 40: 245.9 pg/L
(24.59 pg/dL)
1% PB acetate
group 12 d,
M/F, n = 15
1% PB acetate
group 40 d,
M/F, n = 15
4-115
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Study
Species
(Stock/Strain),
n, Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (pg/dL)
Endpoints Examined
Shvachiv et al.
(2018)
Rat (Wistar) In utero to 28 wk Pregnant Wistar rats were given 18.8 ± 2.0 pg/dL for
Control, M/F,
n = 8
0.2% Pb
acetate,
intermittent Pb
group, M/F,
n = 9
0.2% Pb acetate in drinking
water or tap water. After a 21 d
weaning period, pups were
either continuously exposed to
Pb acetate in drinking water
until 28 wk or were given 8 wk
of Pb abstinence and then
exposed until 28 wk with Pb
acetate
intermittent exposure
24.4 ± 4.9 pg/dL for
continuously exposed
DBP, MAP, Baroreceptor Reflex,
Chemoreceptor Reflex, SBP,
all measured 2-hr post 28-wk
exposure
0.2% Pb
acetate,
permanent Pb
group, M/F,
n = 9
Zhu et al. (2018)
Rat (Sprague
Dawley)
Control
(distilled
water), M,
n = 10
In utero to 1 yr Female rats were given either 0
or 500 mg/L Pb acetate for 10 d
before mating. Male offspring
continued receiving 0 or
500 mg/L Pb acetate for 1 yr
0.28 ± 0.02 mg/L
(28 ± 2 |jg/dL)
SBP, DBP measured post 1 -yr
exposure
0.5 g/L Pb
acetate, M,
n = 10
4-116
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Study
Species
(Stock/Strain),
n, Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (pg/dL)
Endpoints Examined
Zhu etal. (2019)
Rat (Sprague
Dawley)
Control
(distilled
water), M,
n = 100
0.5 g/L Pb
acetate, M,
n = 10
In utero to 1 yr Female rats were given either 0
or 0.5 g/L Pb acetate for 10 d
before mating. Male offspring
continued receiving 0 or 0.5 g/L
Pb acetate for 1 yr
0.27 ± 0.02 mg/L
(27 ± 2 pg/dL)
SBP, DBP measured post 1 -yr
exposure
ACE = angiotensin-converting enzyme; BLL = blood lead level; BW = body weight; d = day(s); DBP = diastolic blood pressure; F = female; M = male; MAP = mean arterial pressure;
NR = not reported; Pb = lead; PND = postnatal day; PP = pulse pressure; SBP = systolic blood pressure; yr = year(s); wk = week(s).
4-117
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Table 4-7
Epidemiologic studies of Pb exposure and coronary and ischemic heart disease
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Jain et al. (2007)
Boston, MA
1991-2001
Cohort
NAS
n = 837
elderly men
(mostly white)
Average
individual
born -1933
Blood (GFAAS with Zeeman
correction) (pg/dL)
Non-cases
Mean (SD): 6.2 (4.3)
Range: 0 to 35
Cases
Mean (SD): 7.0 (3.8)
Range: 1.0 to 20.0
Bone (K-XRF) (pg/g)
Patella
Non-cases
Mean (SD): 30.6 (19.7)
Range: -10 to 165
Cases
Mean (SD): 36.8 (20.8)
Range: 5.0 to 101
Tibia Pb
Non-Cases
Mean (SD): 21.4 (13.6)
Range: -3 to 126
Cases
Mean (SD): 24.2 (15.9)
Range: -5 to 75
IHD
(Ml or angina
pectoris)
Cox proportional hazards models
adjusted forage, BMI, education,
race, smoking status, pack-years
smoked, alcohol intake, history of
diabetes mellitus and hypertension,
family history of hypertension, DBP,
SBP, serum triglycerides, serum
HDL, and total serum cholesterol
BLL >5 pg/dLb
Per 1 SD increase in Pb biomarker
OR over 10-yr follow-up:
1.73 (1.05, 2.87)
Ln (blood Pb)
OR: 1.45 (1.01,
2.06)
Ln (patella Pb)
OR: 2.64 (1.09, 6.37)
Ln (tibia Pb)
OR: 1.84 (0.57, 5.90)
Age at measurement
Mean 67
4-118
-------
Study "oesfg"" Population Exposure Assessment Outcome
Confounders
Effect Estimates and 95% Clsa
Ding etal. (2016)
Boston
1991 through 2011
participants followed
up to 20 yr
Cohort
No minor allele in HFE
rs1799945 (H63D) 1.41 (1.15,
1.73)
rs1800562 (C282Y) 1.36 (1.13,
1.64)
No minor allele HMOX1
rs2071746 1.51 (1.07, 2.14)
rs5995098 1.63 (1.23, 2.14)
At least one minor allele in HMOX1
rs2071749^.b'\ (1.22, 1.86)
No minor allele in APOE
rs429358 1.43(1.17, 1.76)
rs449647 1.29(1.05, 1.60)
rs7412 1.34 (1.10, 1.64)
At least one minor allele in APOE
rs7412 1.53 (1.07, 2.19)
No minor allele in AGT
rs699 2.17 (1.50, 3.12)
rs5046 1.53 (1.27, 1.94)
rs5050 1.36 (1.09, 1.69)
At least one minor allele in AGT
NAS
n = 589
Elderly men
(mostly white)
Average
individual
born -1935
Bone (K-XRF) (pg/g) Mean
(SD)
No CHD
Patella: 29.2 (16.1)
Tibia 20.2 (12.5)
CHD
Patella: 32.1 (18.8)
Tibia: 22.6 (13.5)
Age at measurement:
Mean: 66
Range: 48-96
Incident
Coronary Heart
Disease (Ml,
angina pectoris
or CHD deaths)
Cox regression adjusted for age,
smoking status, BMI, and the ratio
of total cholesterol to
HDL-C level
HR (twofold increase in blood Pb)b
Bone (Patella) 1.36 (1.15, 1.61)
At least one minor allele in VDR
rs1544410 (Bsm1) 1.65 (1.31,
2.08)
rs731236 (Taq1) 1.61 (1.29, 2.02)
rs7975232 (Apa1) 1.28 (1.04, 1.57)
rs1073581 (Fok1) 1.47 (1.17, 1.83)
rs757343 (7>u9V; 1.48 (1.18, 1.85)
At least one minor allele in ALAD
rs1833435 1.11 (0.79, 1.55)
4-119
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Study "oesfg"" Population Exposure Assessment Outcome
Confounders
Effect Estimates and 95% Clsa
Ding etal. (2019)
Boston, MA
United States
August 1991- June
2011
Mean (SE) 8.52
(5.75) yr of follow-up
Cohort
rs50501A1 (1.03, 1.94)
No minor allele in angiotensin II
receptor type 1
rs12695908 1.43 (1.20, 1.74)
No minor allele in glutathione S-
transferase pi 1
rs1695 1.39 (1.10, 1.76)
GRS 1 2.27 (1.50, 3.42)
GRS 2 2.77 (1.78. 4.31)
NAS
n = 594
elderly men
(mostly white)
without CHD
at baseline
Average
individual
born -1935
Bone (K-XRF) (pg/g) Mean
(SD)
CHD
Patella: 32.2 (18.9)
Tibia: 22.6 (13.5)
Non-CHD
Patella: 29.4 (18.9)
Tibia: 20.9 (13.2)
Age at measurement
Mean (SD)
CHD: 65.5 (6.2)
Non-CHD: 66.5 (7.5)
Incident
Coronary Heart
Disease (Ml,
angina pectoris
or CHD deaths)
Cox proportional hazards adjusting
for BMI, total energy intake,
smoking status, TC to HDL ratio,
education level, and occupation
HR (95 % CI) for twofold increase
in Bone Pbb
Patella 1.30 (1.09, 1.56)
Tibia 1.25 (1.06, 1.48)
Prudent Diet
Patella
Low: 1.64 (1.27, 2.11)
High: 1.07 (0.86, 1.34)
Tibia
Low: 1.24 (0.96, 1.59)
High: 1.26 (1.02, 1.55)
Western Diet:
Patella
Low: 1.35 (1.05, 1.72)
High 1.27 (0.96, 1.61)
Tibia
Low: 1.43 (1.14, 1.80)
High: 1.08 (0.86, 1.34)
4-120
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Tonelli etal. (2018) n = 1,278
Plasma Pb (ICP-MS) (pg/dL)
Deciles
Canada
Patients on
incident
hemodialysis
Participant recruited
between March 2005
and November 2012 Average
Cohort (year of
follow-up)
individual
born -1946
0.06
0.19
0.28
0.35
0.44
0.55
0.68
0.83
1.08
10: 1.74
Cardiovascular Logistic regression adjusting for
event (acute Ml, age, sex, race/ethnicity,
percutaneous unemployment prior to dialysis,
coronary year dialysis initiated, dialysis
angioplasty, duration, predialysis care,
coronary artery arteriovenous access, comorbidities
bypass grafting, (AF, Ml, BMI, cancer,
heart failure, cerebrovascular disease, CHF, lung
and stroke or disease, diabetes, dementia,
transient hypertension, liver disease, PVD,
ischemic attack) psychiatric disease, substance
misuse), albumin, and creatinine.
*AII variables were considered
candidate variables and were
included based on stepwise
regression results
ORb
Cardiovascular events: NR
Authors indicate a null relationship
between blood Pb deciles and all-
cardiovascular events, results not
reported
Cho et al. (2016)
KNHANES
Blood (GFAAS with Zeeman
>10% increase Logistic regression adjusted for
ORb (95%
CI):
n = 5,361
correction) (pg/dL)
in 10-yrCHD BMI, triglycerides, and LDL-C
Males
South Korea
Participants in
KNHANES
Males
Mean (SE): 2.81 (0.32)
Risk (FRS)
based on age,
gender, SBP,
Q2 vs. Q1
Q3 vs. Q1
1.59 (1.03, 2.46)
2.31 (1.52, 3.50)
2008-2019
aged 20-70 y
Q1: 0.71-2.13
total cholesterol,
Q4 vs. Q1
3.13 (2.9, 4.69)
Q2: 2.13-2.70
and HDL-C)
Females
Cross-sectional
Average
Q3: 2.70-3.52
Q2 vs. Q1
1.84 (0.61, 5.55)
individual
Q4: 3.52-26.51
Q3 vs. Q1
1.43 (0.44, 4.59)
born -1973
Females
Q4 vs. Q1
0.88 (0.26, 2.97)
Mean (SE): 2.04 (0.02)
Q1
Q2
Q3
Q4
0.42-1.49
1.49-1.95
1.95-2.51
2.51-9.59
Age at measurement (years):
Mean (SE)
Males: 39.3 (0.30)
4-121
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Females: 40.9 (0.30)
Choi etal. (2020)
KNHANES
Blood (GFAAS with Zeeman
10-yr
Multiple linear regression analysis
Linear increase in 10-yr ASCVD
n = 2,424
correction) (pg/dL)
atherosclerotic
adjusting for age, income, job type,
risk score (Q4 vs. Q1 )b c
Korea
cardiovascular
physical activity, location, and sleep
Participants in
Pb distribution NR
disease
Males: 0.117 (0.01, 0.23)
2016-2017
KNHANES
aged 40-80 yr
(ASCVD) risk
Urban: 0.133 (0.01, 0.25)
Rural: 0.079 (-0.15, 0.31)
Cross-sectional
Average
individual
born -1956
<7 hr sleep: 0.097 (-0.03, 0.23)
>7 hr sleep: 0.183 (0.01, 0.36)
Females: 0.072 (-0.00, 0.15)
Urban: 0.038 (-0.05, 0.12)
Rural: 0.212 (0.05, 0.38)
<7 hr sleep: 0.110 (0.016, 0.20)
>7 hr sleep: 0.021 (-0.09, 0.13)
Park and Han (2021)
KNHANES
Blood Pb (GFAAS with
10%-20% and
Logistic regression adjusted for
OR (<10% increase in 10-yr CVD
n = 1,929
Zeeman correction (pg/dL)
>20% increase
SBP, HDL cholesterol, and total
risk as referent)bcd
South Korea
Distribution: NR
in 10-yr CVD
risk estimated
cholesterol
Males
>20 y
using FRS
10%-20% 10 yr CVD risk (vs.
KNHANES VII-1
<10%): 2.407 (1.885, 3.075)
(2017)
Average
>20% 10 yr CVD risk (vs. <10%):
2.847 (2.020, 4.011)
individual
Cross-sectional
born in or
before -1997
Females
10%-20% 10 yr CVD risk (vs.
<10%): 1.051 (0.676, 1.633)
>20% 10 yr CVD risk (vs. <10%):
0.706 (0.188, 2.659)
4-122
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Nguyen et al. (2021)
South Korea
KNHANES IV (2009),
V (2010-2012), VI
(2013), and VII
(2016-2017)
Cross-sectional
KNHANES
n = 9,602
>20 y
Average
individual
born -1965
Serum Pb (GFAAS with
Zeeman correction (pg/dL)
GM (95% CI)
2.02 (2.00-2.03)
Age at measurement
Mean (SD)
Males: 47.76 (15.25)
Females: 46.87 (15.16)
10-yr CVD risk Multivariable models adjusted for
estimated using serum cotinine, age group, sex,
FRS high-risk drinking, physical activity,
BMI, family history of CVDs,
diabetes or dyslipidemia, and type
2 diabetes.
*Assume linear regression, but not
specified
Linear increase in 10-yr CVD risk
score (log2 transformed blood Pb)b
0.104 (0.016, 0.214)
Ochoa-Martinez et al. Women living Blood Pb (GFAAS) (pg/dL)
(2018)
San Luis Potosi
Mexico
2015-2016
Cross-sectional
in
communities
with a high-
risk of
environmental
Pb
contamination
n = 175
Average
individual
born -1967
mean (SD) 11.5 (9.0)
Tertiles
T1
T2
T3
<3.5
3.6-9.0
>9.1
Age at measurement:
mean (SD): 48.5 (18.0)
Predictive CVD
biomarkers
[asymmetric
dimethylarginine
(ADMA),
FABP4,
adiponectin,
and chemerin]
Linear regression controlling for
age, weight, waist circumference,
hip circumference, SBP, DBP, BMI,
body fat %, visceral fat %, glucose,
triglycerides, total cholesterol, HDL-
C, LDL-C
Predictive CVD biomarkersb
ADMA (pmol/L):
T2: 0.51 (-0.25, 0.69)
T3: 0.75 (0.15, 1.85)
FABP4 (ng/mL):
T2: 11.0 (-15.0, 16.0)
T3: 27.5 (10.0, 34.5)
Adiponectin (|jg/mL):
T2: 9.50 (-17.0, 21.0)
T3: 12.5 (-7.5-, 18.0)
Chemerin (ng/mL):
T2: 195 (-75.0, 275)
T3: 220 (-25, 300)
4-123
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Wan etal. (2021)
China
May-August 2018
Cross-sectional
Environmental Blood (AAS) (pg/dL)
Pollutant
Exposure and
Metabolic
Diseases in
Shanghai
(METAL
study)
n = 4,234
Median (IQR):
2.6 (1.8-3.6)
Age at measurement
Median (IQR):
67 (62-72) yr
Presence of
CVD (Self-
reported
diagnosis by a
physician,
including CHD,
Ml or stroke)
Linear or logistic regression
adjusting for age, sex, duration of
diabetes, education status, current
smoking, BMI, HbA1c,
dyslipidemia, hypertension
OR (95% CI) <4th vs 1 st quartile of
Blood Pb)b
1.44 (1.17, 1.76)
Average
individual
born -1951
AAS = atomic absorption spectrometry; ADMA = asymmetric dimethylarginine; AF = atrial fibrillation; AGT = angiotensinogen; ALAD = 6-aminolevulinic acid dehydratase;
APOE = apolipoprotein E; ASCVD = atherosclerotic cardiovascular disease; BLL = blood lead level; BMI = body mass index; C282Y HFE = mutant of the HFE wildtype;
CHD = congenital heart disease; CHF = congestive heart failure; CI = confidence interval; CVD = cardiovascular disease; DBP = diastolic blood pressure; FABP4 = adipocyte fatty
acid-binding protein 4; FRS = Framingham risk score; GFAAS = graphite furnace atomic absorption spectrometry; GRS = genetic risk score; HbA1c = hemoglobin A1c; HDL-
C = high-density lipoprotein cholesterol; HFE = hemochromatosis gene; HMOX1 = heme oxygenase-1; HR = hazard ratio; ICP-MS = inductively coupled plasma mass spectrometry;
IHD = ischemic heart disease; IQR = interquartile range; KNHANES = Korea National Health and Nutrition Examination Survey; K-XRF = K-shell X-ray fluorescence; LDL-C = low-
density lipoprotein cholesterol; METAL = Environmental Pollutant Exposure and Metabolic Diseases in Shanghai ; Ml = myocardial infarction; NR = not reported; OR = odds ratio;
Pb = lead; PVD = peripheral vascular disease; Q = quartile; SBP = systolic blood pressure; SD = standard deviation; SE = standard error; T# = fertile #; TC = total cholesterol;
VDR = vitamin D receptor; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th~90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bUnable to be standardized.
Confidence intervals estimated from standard error.
increment unclear.
4-124
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Table 4-8
Epidemiologic studies of Pb exposure and cardiac function
Study DCesfgnnd Population Exposure Assessment Outcome
Schwartz (1991)
United States
1976-1980
Cross-sectional
Confounders Effect Estimates and 95% Clsa
Logistic regression adjusting OR
for age, sex, and race 1 028 (1.009, 1.048)
NHANES II
n = 9,932
Average
individual born
-1931
Blood- Method NR
Distribution NR
Left ventricle
hypertrophy
(based on
body habitus,
BMI, and tricep
skinfold)
Yang etal. (2017)
Belgium
baseline 1985-
2000; follow-up:
2005-2010
Median follow-up
11.9 yr
Cohort
Cadmium in
Belgium study
n = 179
Average
individual born
-1953
Blood (ETAAS with
Zeeman correction) (pg/dL)
GM 4.14
Age at measurement
Mean: 39.1
Left ventricle Linear regression adjusting
structure and for measures at the time of
function the echocardiography
including age, sex, MAP,
heart rate, BMI, fasting
plasma glucose, total to HDL
cholesterol ratio, serum
creatinine, y-
glutamyltransferase,
smoking, and
antihypertensive medication
class
For each doubling of blood Pbbc
LV Structure
LVMI, g/m2-1.399 (-4.504,1.707)
End diastolic diameter, cm -0.064 (-0.134,
-0.006)
RWT 0.0065 (-0.0031, 0.0162)
Systolic LV Function
Ejection fraction, % 0.190 (-1.293, 1.675)
GLS, % -0.497 (-0.957, -0.038)
RLS, % -0.784 (-1.482, -0.087)
RLS rate, (s-1) -0.071 (-0.124, -0.019)
RRS, % -2.316 (-4.748, -0.115)
RRS rate, (s-1) -0.135 (-0.292, 0.022)
Diastolic LV Function
E peak, cm/s 1.308 (-1.120, 3.736)
E/A ratio -0.036 (-0.085, 0.014)
e' peak, cm/s -0.188 (-0.494, 0.118 (E/e'
ratio 0.172 (-0.133, 0.477)
4-125
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Lindetal. (2012)
Uppsala, Sweden
Cross-sectional
Prospective
Investigation of
the Vasculature
in Uppsala
Seniors (PIVUS)
study
Blood (ICP-MS) (pg/dL)
Median (IQR): 1.72 (1.22,
2.28)
Left ventricle Linear regression adjusting
structure for sex, BP, antihypertensive
medication, diabetes, and
BMI
Per In-transformed unit increase in serum
Pbb
LVMI, g/m2 -0.7-3 -(2.20, 0.74)
RWT 0.011 (-0.001, 0.022)
n = 993
Elderly (70 yr)
individuals
Chen etal. (2021) n = 486
Guangdong
province
China
2018
Cross-sectional
Preschool
children (aged
2-6) from two
towns with
similar SES but
different Pb
exposure
Average
individual born
-2013
Blood (GFAAS) (pg/dL)
Median (IQR):
Exposed: 4.51 (3.70-5.67)
Reference: 3.98 (3.25-
4.84)
Age at measurement
Mean (SD):
Exposed: 4.74 (0.84)
Reference: 4.75 (1.01)
Left ventricle Linear regression adjusted
structure and for gender, age, BMI, e-
function waste contamination w/ in
50 m of residence, residence
as workplace, distance of
residence from road, family
member daily smoking,
monthly household income,
maternal work associated
with e-waste, duration of
outdoor play, child contact
with e-waste, washing hands
before eating, nail biting
habit, chewing pencil habit,
yearly canned food
consumption, yearly
fruit/vegetable consumption,
yearly iron rich food
consumption, yearly marine
product consumption, and
yearly salted food
consumption
Ln-transformed parameters per one-unit
increase in blood Pbb
-0.007, 0.001)
mm -0.001 (-0.003,
IVS, cm -0.004 (
LV posterior wall,
0.001)
Ejection fraction, % 0.001 (-0.002, 0.001)
4-126
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Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Gump et al. (2005) Oswego
Oswego, NY (born
at a single hospital
in New York from
1991-94)
Cohort
Children's Study
n = 122 children
aged 9.5 yr
Average
individual born
-1990
Cord blood (ETAAS)
(Hg/dL)
Mean (SD): 2.97 (1.75)
Child blood (ETAAS and
ASV) (pg/dL)
Mean (SD): 4.62 (2.51)
Age of child blood Pb
measurement
Mean 2.6
Stroke volume, Multivariate linear regression Cord BLL
cardiac output
adjusted for HOME score,
SES, birth weight, child BMI,
child sex
No association, results not reported
Childhood BLL
Stroke volume, mL -0.069 (-0.124, -0.015)
Cardiac output, L/min -0.056 (-0.113,
0.001)
Gump et al. (2011) n = 140 children Blood (ICP-MS) (pg/dL)
ages 9-11 yr
Oswego, NY
Cross-sectional
GM
1.01
Q1:
0.14-0.68
Q2:
0.69-0.93
Q3:
0.94-1.20
Q4:
1.21-3.76
Stroke volume, Linear regression models
cardiac output adjusted for sex, SES, BMI,
and age
Change in Stroke Volume (%) across
quartiles in response to acute stressbd
Q1: 2.23, Q2: 0.91, Q3: -3.47, Q4: -0.89,
p for trend = 0.04
Change in Cardiac Output (%) across
quartiles in response to acute stressbd
Q1: 3.26, Q2: 1.19, Q3: -2.31, Q4: -0.20,
p for trend = 0.05
A = peak late diastolic velocity; ASV = anodic stripping voltammetry; BLL = blood lead level; BMI = body mass index; BP = blood pressure; E = peak early diastolic velocity; e' = peak
early diastolic mitral annular velocity; ETAAS = electrothermal atomic absorption spectrometry; GFAAS = graphite furnace atomic absorption spectrometry; GLS = global longitudinal
strain; GM = geometric mean; GSD = geometric standard deviation; HDL = high-density lipoprotein; HOME = Health Outcomes and Measures of the Environment; ICP-
MS = inductively coupled plasma mass spectrometry; IQR = interquartile range; IVS = interventricular septum; LV = left ventricular; LVMI = left ventricular mass index; MAP = mean
arterial pressure; NHANES = National Health and Nutrition Examination Survey; NR = not reported; OR = odds ratio; Pb = lead; PIVUS = Prospective Investigation of the Vasculature
in Uppsala Seniors; RLS = regional longitudinal strain; RRS = regional radial strain; RWT = relative wall thickness; SD = standard deviation; SES = socioeconomic status;
Q = quartile.
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bUnable to be standardized.
¦Corrected for regression dilution bias using quintile method.
Confidence intervals not provided and unable to calculate based on given information.
4-127
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Table 4-9 Animal toxicological studies of cardiac function
g^ucjy Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (pg/dL)c
Endpoints
Examined
Wildemann et al. (2015) Rat (Wistar)
Control (tap water), M,
n = 6
Unknown start age
for 4 wk
Tap water with 0.2% nitic
acid, or 357 or 1607 pg/kg
BW/d Pb acetate in
drinking water for 4 wk
1.4 ± 1.2 pg/L for tap water/
0.2% nitic acid
(0.14 ± 0.12 pg/dL)
Stroke volume and
cardiac output post
4-wk exposure
357 or 1607 pg/kg BW/d,
M, n = 5 per group
17 ± 7 pg/L for 357 pg/kg
BW/d Pb acetate
(1.7 ± 0.7 pg/dL)
86 ±29 pg/L for 1607 pg/kg
BW/d Pb acetate
(8.6 ±2.9 pg/dL)
Silvaetal. (2015) Rat (Wistar)
Control (distilled water)
M, n = 6
3 mo old rats
exposed for 15 d
100 ppm Pb acetate in
drinking water for 15 d
12.3 ± 2 pg/dL
Force generation in
LV papillary muscle
following pulse
stimulation post 15-d
exposure
100 ppm Pb acetate group
M, n = 6
Time to peak tension
and 90% relaxation
post 15-d exposure
Inotropic force
following calcium or
isoproterenol
stimulation post 15-d
exposure
4-128
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Study
Species (Stock/Strain), Timing of
n, Sex Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (|jg/dL)c
Endpoints
Examined
Fioresi et al. (2014)
Rat (Wistar)
Control (tap water), M,
n = 9-12
100 ppm group, M, n = 9-
12
Age 2 mo to 3 mo
100 ppm Pb acetate in
drinking water for 30 d
<0.5 |jg/dL for control
13.6 ± 1.07 |jg/dL for
100 ppm group
LVSP and RVSP, left
and right diastolic
pressure all
measured post 30 d
exposure
Isometric contraction
force, time to peak
contraction, and
relaxation rates in LV
papillary muscle post
30 d exposure
Contractile force
following calcium
treatment in LV
papillary muscle 30 d
post exposure
BW = body weight; d = day(s); LV = left ventricular; LVSP = right ventricular systolic pressure; M = male; mo = month(s); Pb = lead; RVSP = right ventricular systolic pressure;
wk = week(s).
4-129
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Table 4-10 Animal toxicological studies of Pb exposure and endothelial dysfunction
Study
Species (Stock/Strain), n, Timing of Exposure Details BLL as Reported Endpoints
Sex Exposure (Concentration, Duration) (Mg'dL) Examined
Gaspar and Cordellini
(2014)
Rat (Wistar)
Rat (Wistar)
Control (tap water), M, n :
500 ppm group, M, n = 6
In utero to PND 22
Pregnancy day 0: females
divided into tap water and
500 ppm Pb acetate in drinking
water groups. Exposure lasted
through pregnancy. At birth,
pups were exposed to Pb (or
control) through nursing. Pups
were weaned at 22 d.
<5 |jg/dL at all time
points for tap water
19.98 ± 6.31 -
PND 52
13.15 ± 0.97-
PND 70
Vascular reactivity
in aortic rings post
exposure at
PND 23, 52, 70,
and 100
11.17 ± 2.11 -
PND 100
Nunes et al. (2015)
Rat (Wistar)
Control (distilled water), M,
n = 16
2-mo-old rats
exposed for 30 d
100 ppm Pb acetate in drinking
water for 30 d.
NR for control
8.4 |jg/dL for
100 ppm
Vascular reactivity
in aortic rings
measured post 30-
d exposure
100 ppm Pb acetate
treatment, M, n = 16
5-8 rats from control or
100 ppm group for other
treatments
(e.g., phenylephrine in control
or Pb-treated mice)
BLL = blood lead level; d = day(s); M = male; NR = not reported; Pb = lead; PND = postnatal day.
4-130
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Table 4-11
Epidemiologic studies of Pb exposure cardio electrophysiology and arrythmia
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Chena et al.
(1998)
Boston, MA
NAS August 1991
and December
1995
NAS
n = 750
Elderly men (mostly
white)
Average individual
born -1925
Blood Pb (GFAAS with
Zeeman correction)
(pg/dL)
Mean (SD): 5.79 (3.44)
Bone (K-XRF) (pg/g)
Mean (SD):
Patella: 30.82 (19.19)
ECG
conduction
(QTc, QRSc)
Linear regression adjusted for
age and DBP. Model for QTc
additionally adjusted for
alcohol intake and BMI. Model
for QRSc additionally adjusted
for fasting glucose level.
QTc (msec)
Bone (Patella) Pb
<65 yr: 3.00 (0.16, 5.84)
>65 yr: 0.39 (-1.05, 1.83)
Tibia Pb
<65 yr: 5.03 (0.83, 9.22)
>65 yr: 1.41 (-0.67, 3.49
Cross-sectional
Tibia: 22.9 (13.36)
Age at measurement
Mean (SD): 67.81 (7.27)
QRSc (msec)
Patella Pb
<65 yr: 2.23 (0.10, 4.36)
>65 yr: -0.11 (-1.07, 0.85)
Tibia Pb
<65 yr: 4.83 (1.83, 7.83)
>65 yr: -0.83 (-2.21, 0.56)
No association with blood Pb
Eum et al. (2011)
Boston, MA
NAS
n = 600
Elderly men (mostly
white)
Blood (GFAAS with
Zeeman correction)
(pg/dL)
Mean (SD): 5.8 (3.6)
ECG
conduction
(QTc, QRSc)
Linear regression adjusted for
age, education, smoking, BMI,
albumin-adjusted serum Ca2+,
and diabetes status at
baseline, years between ECG
Bone (Tibia) Pbb
Adjusted 8-yr change
QTc(Q1 reference)
Q1: 7.49 (1.22, 13.75) msec
NAS 1989 and
1996
Average individual
born -1925
Bone (K-XRF) (pg/g)
Mean (SD):
tests, and QT-prolongation
drugs at the time of ECG
measurement.
Q3: 7.94 (1.42, 14.45) msec
p for trend = 0.03
Cohort
Patella: 30.3 (17.7)
Tibia: 21.6 (12.0)
Tibia Quartiles:
Q1: <16
Q2: 16.0-23
QRSc (Q1 reference)
Q2: 0.52 (-3.60, 4.65) msec
Q3: 5.94 (1.66, 10.22) msec
p for trend = 0.005
Q3: >23
4-131
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Age at measurement:
Mean 67
No associations with patella
or blood Pb
Park et al. (2009)
Boston, MA
NAS August 1991
and December
1995
Cross-sectional
NAS
n = 613
Elderly men (mostly
white)
Average individual
born -1925
Blood (GFAAS with
Zeeman correction)
(Mg/dL)
Median (IQR): 5 (4-7)
Bone (K-XRF) (pg/g)
Median (IQR)
Patella: 26 (18-37)
Tibia: 19 (14-27)
QTc interval Linear regression models
adjusted forage, BMI, smoking
status, serum Ca2+, and
diabetes. No SES indicator
was considered.
QTc interval (msec)
0.433 (-0.253, 1.12)
Patella Pb:
1.389 (0.068, 2.711)
Tibia Pb:
2.192 (0.227, 4.158)
Age at measurement
Mean 67
Jina et al. (2019)
United States
NHANES III
1988-1994
Cross-sectional
NHANES
n = 7,179
Participants without
self-reported history of
Ml (or ECG results
indicating Ml), without
a history of CHF
Average individual
born -1934, -1936,
-1932, -1935
Blood (GFAAS) (pg/dL)
GM
Men: 4.10
Women: 2.93
Age at measurement:
Mean (SD)
Men
T1
T2
T3
57.12 (0.51)
55.42 (0.53)
57.00 (0.65)
Women
T1
T2
T3
59.01 (0.64)
55.88 (0.62)
59.45 (0.98)
Ventricular
arrhythmia
(QRS-T angle)
Spatial QRS-T
angle
estimated
using a 12-Pb
ECG.
Multivariate weighted logistic
regression adjusting for
impaired fasting glucose,
hypertension, poverty index,
age, race, and smoking status.
One-unit increase in log of
blood Pbb
OR (95% CI) (3rd vs. 1st
fertile)
Men: 1.35 (1.05, 1.74)
Women: 1.05 (0.82, 1.36)
4-132
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Park et al. (2006) NAS
n = 413
Boston, MA
2000-2004
Cross-sectional
Elderly men (mostly
white)
Average individual
born -1935
Bone (K-XRF) (pg/g)
Median (IQR)
Tibia: 19 (11-28)
Patella (measured within
6 mo of HRV):
23 (15-34)
Estimated Patella
(accounting for time
difference):
16.3 (10.4-25.8)
Age at measurement:
Mean: 67
HRV
Linear regression models
adjusted for age, cigarette
smoking, alcohol consumption,
room temperature, season
(Model 2) BMI, fasting blood
glucose, HDL-C, triglyceride,
use of p-blockers, Ca2+
channel blockers, and/or ACE
inhibitors. No SES indicator
was considered.
HRV
Tibia
HF: -0.529 (-2.265, 1.206)
nu
LF: 0.529 (-1.206, 2.265) nu
Log LF/HF:
1.941 (-6.941, 10.824) %
Corrected Patella
HF: -0.39 (-2.013, 1.234) nu
LF: 0.39 (-1.234, 2.013) nu
Log LF/HF:
1.948 (-6.136, 10.032)
Effect estimates were more
pronounced among those
with greater# metabolic
abnormalities.
Gump et al.
(2011)
Oswego, NY
Cross-sectional
n = 140
Children aged 9-11 yr
Blood (ICP-MS) (pg/dL)
GM: 1.01
Heart rate Linear regression adjusted for
sex, SES, BMI, and age.
Q1
Q2
Q3
Q4
0.14-0.68
0.69-0.93
0.94-1.20
1.21-3.76
Change in heart rate
(beats/min) across quartiles
in response to acute stressbc
Q1: 0.91, Q2: 0.19, Q3: 0.86,
Q4: 0.58, p for trend = 0.85
No association between Pb
levels and baseline
cardiovascular levels
Gump et al.
(2017)
New York
Environmental
Exposures and Child
Health Outcomes
study
n = 203
Children aged 9-11
Blood (ICP-MS) (pg/dL) Heart rate
Mean (SD): 0.98 (0.61) variability
Range: 0.19-3.25
Linear regression adjusted for
sex, race, age, and SES.
No association between BLLs
and HRV. Results not shown
(p > 0.25)
4-133
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Environmental
Exposures and
Child Health
Outcomes
Halabickv et al.
(2022)
Jintan, China
Children aged 3-5
in 2004-2005
wave, aged 11-12
in 2011-2013
wave
Cohort
China Jintan Child
Cohort Study
n = 408
Average individual
born -2000
Whole blood (GFAAS)
(pg/dL)
Median (IQR)
3-5 yr: 6.4 (4.9-8.0)
11-13 yr: 2.9(2.3-3.6)
Age at measurement:
3-5 and 11-13 yr
HRV in
children during
a stress test
Age at
outcome -12
General linear models
adjusted for parental
occupation, child sex, serum
Fe, crowded neighborhood
*No baseline HRV
measurement and no
adjustment for BMI or physical
activity. Serum Fe less precise
measure of iron status
opposed to ferritin or
transferrin.
Change in planning phase
HRV per log-transformed
increase in blood Pbb
3-5 yr Pb 0.03 (-0.02, 0.09)
12 yr Pb -0.04 (-0.16, 0.07)
Change in speaking phase
HRV per log-transformed
increase in blood Pbb
3-5 yr Pb: 0.06 (0.01, 0.12)
12 yr Pb: -0.05 (-0.18, 0.08)
ACE = angiotensin-converting enzyme; BLL = blood lead level; BMI = body mass index; Ca2+ = calcium ion(s); CHF = congestive heart failure; CI = confidence interval;
DBP = diastolic blood pressure; ECG = electrocardiogram; GFAAS = graphite furnace atomic absorption spectrometry; GM = geometric mean; HDL = high-density lipoprotein;
HF = high-frequency power in normalized units; HRV = heart rate variability; ICP-MS = inductively coupled plasma mass spectrometry; IQR = interquartile range; K-XRF = K-shell
X-ray fluorescence; LF = low-frequency; Ml = myocardial infarction; mo = month(s); NAS = Normative Aging Study; NHANES = National Health and Nutrition Examination Survey;
nu = normalized units; OR = odds ratio; Pb = lead; Q = quartile; QRSc = corrected QRS duration; QTc = corrected QT interval; SD = standard deviation; SES = socioeconomic
status; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bUnable to be standardized.
Confidence intervals not provided and unable to calculate based on given information
4-134
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Table 4-12 Animal toxicological studies of Pb exposure and cardiac electrophysiology
Study (Stock/Strata), n, ™"9°< Exposure Details BLL as Reported Endpolnts Examlned
' v Sex ' Exposure (Concentration, Duration) (Mg'dL) ^
Fioresi et al. (2014)
Rat (Wistar)
Control (tap water),
M, n =9-12
100 ppm group, M,
n = 9-12
Age 2 mo to
3 mo
100 ppm Pb acetate in drinking water for 30 d
<0.5 |jg/dL for
control
13.6 ± 1.07 |jg/dL
for 100 ppm group
Heart rate post 30-d
exposure
Wildemann et al.
(2015)
Rat (Wistar)
Control (tap water),
M, n = 6
357 or 1607 |jg/kg
BW/d, M, n = 5 per
group
Unknown start Tap water with 0.2% nitic acid, or 357 or
age for 4 wk 1607 |jg/kg BW/d Pb acetate in drinking water for
4 wk
1.4 ± 1.2 |jg/L for
tap water/ 0.2%
nitic acid
(0.14 ± 0.12 pg/dL)
17 ± 7 pg/L for
357 pg/kg BW/d
Pb acetate
(1.7 ± 0.7 pg/dL)
Heart rate measured
baseline and post 4-wk
exposure
PR and QRS interval
post 4-wk exposure
86 ± 29 pg/L for
1607 pg/kg BW/d
Pb acetate
(8.6 ± 2.9 pg/dL)
4-135
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Species
Study (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported
(Hg/dL)
Endpoints Examined
Shvachiv et al. (2018)
Rat (Wistar)
Control, M/F, n = f
0.2% Pb acetate,
intermittent Pb
group, M/F, n = 9
In utero to Pregnant Wistar rats were given 0.2% Pb acetate
28 wk in drinking water or tap water. After a 21-d
weaning period, pups were either continuously
exposed to Pb acetate in drinking water until 28 wk
or were given 8 wk of Pb abstinence and then
exposed until 28 wk with Pb acetate
18.8 ±2.0 |jg/dL
for intermittent
exposure
24.4 ± 4.9 |jg/dL
for continuously
exposed
Heart rate, HF, LF,
LF/HF measured 24-hr
post 28-wk exposure
0.2% Pb acetate,
permanent Pb
group, M/F, n = 9
Zhu et al. (2018) Rat (Sprague In utero to 1 yr Female rats were given either 0 or 500 mg/L Pb 0.28 ± .02 mg/L Heart rate, HF, LF,
Dawley) acetate for 10 d before mating. Male offspring (28 ± 2 |jg/dL) LF/HF measured post 1-
Control (distilled continued receiving 0 or 500 mg/L Pb acetate for yr exposure
water), M, n = 100 1 V
0.5 g/L Pb acetate,
M, n = 10
M,n = 100
0.5 g/L Pb acetate,
M, n = 10
Zhu et al. (2019) Rat (Sprague In utero to 1 yr Female rats were given either 0 or 0.5 g/L Pb 0.27 ± 0.02 mg/L HF, LF, LF/HF, heart
Dawley) acetate for 10 d before mating. Male offspring (27 ± 2 |jg/dL) measured post 1-yr
Control (distilled continued receiving 0 or 0.5 g/L Pb acetate for exposure
water), M, n = 100 1 V
0.5 g/L Pb acetate,
M, n = 10
BLL = blood lead level; BW = body weight; d = day(s); F = female; HF = high-frequency; LF = low-frequency; M = male; mo = month(s); Pb = lead; PR = prevalence ratio;
wk = week(s); yr = year(s).
4-136
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Table 4-13
Epidemiologic studies of Pb exposure and atherosclerosis and peripheral artery disease
Reference and
Study Design
Study
Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Wan et al. (2021)
METAL study
Blood (AAS) (pg/dL)
CCA plaques and
Linear or logistic
OR (4th vs. 1st quartile of Blood Pb)b
n = 4,234
Median (IQR)
diameter
regression adjusting for
1.53 (1.29, 1.82)
China
2.6 (1.8-3.6)
age, sex, duration of
diabetes, education
May-August 2018
Average
status, current smoking,
individual born
Age at measurement
BMI, HbA1c,
Cross-sectional
-1951
Median (IQR):
67 (62-72) yr
dyslipidemia,
hypertension
Yu et al. (2020)
Belgium
Cadmium in
Belgium
n = 267
Blood Pb collected in Average
1985-2005, arterial
individual born
stiffness measured a ~1958
median of 9 yr later
Cohort
Blood (ETAAS) (pg/dL)
GM (IQR):
2.93 (1.8-4.7)
Age at measurement:
Mean 37 yr
Arterial stiffness
Hemodynamic
measures
Linear multivariable
models adjusting for sex,
enrollment characteristics
(age, BMI, smoking,
drinking, serum total to
HDL-C ration, plasma
glucose, eGFR
(estimated from serum
creatinine), SES), the
time interval between
measurement of
exposure biomarkers and
hemodynamic
assessment, and
antihypertensive drug
treatment at enrollment
and follow-up
Time-dependent hemodynamics per doubling of
Pb concentration15
Augmentation ratio, % 1.74 (0.95, 2.53)
Augmentation index, % 3.03 (1.56, 4.50)
Pressure amplification -0.06 (-0.08, -0.04)
Pulse wave velocity, m/s 0.14 (-0.08, 0.35)
Forward pulse peak time, ms 6.62 (2.21, 11.0)
Backward pulse peak time, ms 1.02 (-1.31, 3.35)
Forward PP amplitude, mmHg
-0.43 (-1.92, 1.06)
Backward PP amplitude, mmHg
1.02 (0.02, 2.02)
Reflection index, %: 3.98 (1.71, 6.24)
4-137
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Reference and
Study Design
Study
Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Kimetal. (2021)
South Korea
2011-2018
Cross-sectional
n =2,193
Adults >2 yr of
age who
completed
voluntary
medical
examinations at
the Chonnam
National
University
Hwasun
Hospital
Average
individual born
-1961
Whole blood Pb
(GFAAS) (|jg/dL)
Mean (Median)
All participants:
2.71 (2.53)
Men: 2.98 (2.78)
Women: 2.18 (2.03)
Age at measurement:
Mean (SD): 53.5 (8.3)
Range: 23-81
Moderate to
severe CAS
(>25% stenosis)
Logistic regression
adjusted for age, sex,
hypertension, diabetes
mellitus, dyslipidemia,
BMI, regular exercise,
smoking, and alcohol
drinking
OR
All participants: 1.14 (1.02, 1.26)
Men: 1.13 (1.01, 1.27)
Women: 1.10 (0.86, 1.41)
Qin etal. (2021)
United States
NHANES 2013-2014
Cross-sectional
NHANES
n = 1,503
>40 yr
Average
individual born
-1960
Blood (ICP-MS)
(|jg/dl_)c
Mean (SD): 1.45 (1.31)
Q1: 0.16-0.80
Q2: 0.81-1.21
Q3: 1.22-1.84
Q4: 1.85-24.6
Age at measurement
Mean: 52.7
AAC score (0-24 Linear or logistic
for total score),
and severe AAC
(AAC score >6)
regression adjusted for
sex, age, race,
education, BMI, BP,
creatinine, A1c, uric acid,
serum calcium, serum
phosphorus, total
cholesterol, cotinine,
hemoglobin, hydroxy-
vitamin D, hypertension,
diabetes
Change in AAC score
Per |jg/dL increase in blood Pb 0.15 (0.02, 0.27)
Q2 vs. Q1: 0.58 (0.15, 1.02)
Q3 vs. Q1: 0.60 (0.14, 1.07)
Q4 vs. Q1: 0.99 (0.50, 1.48)
OR
Per |jg/dL increase in blood Pb 1.11 (1.00, 1.22)
Q2 vs. Q1
Q3 vs. Q1
Q4 vs. Q1
1.68 (0.86, 3.25)
2.15 (1.10, 4.19)
3.72 (1.94, 7.12)
4-138
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Reference and
Study Design
Study
Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Muntner et al. (2005)
United States
NHANES 1999-2002
Cross-sectional
NHANES
n = 9,961
Average
individual born
-1954
Or
Average
individual born
in or before
-1983
Blood (GFAAS)
(Hg/dL)
Mean (95% CI): 1.64
(1.59-1.68)
PAD
Q1
Q2
Q3
Q4
<1.06
1.06-1.63
1.63-2.47
>2.47
Age at measurement
Mean NR
Logistic regression
models adjusted for age,
race/ethnicity, sex,
diabetes mellitus, BMI,
cigarette smoking,
alcohol consumption,
high school education,
health insurance status
OR (95% CI) (vs. 1st quartile)b
Q2
Q3
Q4
1.00 (0.45, 2.22)
1.21 (0.66, 2.23)
1.92 (1.02, 3.61)
Navas-Acien et al.
(2004)
United States
1999-2000
Cross-sectional
NHANES
n =2,125
participants,
age >40 yr,
Average
individual born
in or before
-1960
Blood (GFAAS)
(Hg/dL)
Q1
Q2
Q3
Q4
<1.45
1.45-2.07
2.07-2.90
>2.90
PAD
Logistic regression
adjusted for age, sex,
race, education, BMI,
alcohol intake,
hypertension, diabetes,
hypercholesterolemia,
eGFR, C-reactive
protein, self-reported
smoking status, serum
cotinine and Cd
OR (95% CI) (vs. 1st quartile)b
Q2
Q3
Q4
1.63 (0.50-5.27)
1.77 (0.55-5.63)
2.52 (0.75-8.51)
AAC = abdominal aortic calcification; AAS = atomic absorption spectrometry; BMI = body mass index; BP = blood pressure; CAS = coronary artery stenosis; CCA = common carotid
artery; Cd = cadmium; CI = confidence interval; eGFR = estimated glomerular filtration rate; ETAAS = electrothermal atomic absorption spectrometry; GFAAS = graphite furnace
atomic absorption spectrometry; GM = geometric mean; HbA1c = hemoglobin A1c; HDL-C = high-density lipoprotein cholesterol; ICP-MS = inductively coupled plasma mass
spectrometry; IQR = interquartile range; METAL = Environmental Pollutant Exposure and Metabolic Diseases in Shanghai; mo = month(s); NHANES = National Health and Nutrition
Examination Survey; NR = not reported; OR = odds ratio; PAD = peripheral artery disease; Pb = lead; PP = pulse pressure; Q = quartile; SD = standard deviation;
SES = socioeconomic status; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bUnable to be standardized.
°Results reported as ng/dL but assumed to be ug/dL based on data source (NHANES).
4-139
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Table 4-14
Animal toxicological studies of Pb exposure and atherosclerosis
Study
Species (Stock/Strain),
n, Sex
Timing of Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
Xu et al. (2015)
Rat (Sprague Dawley)
Control 1, M/F, n = 6
1% PB acetate group
12 d, M/F, n = 15
6-7-wk old rats exposed for
12 or 40 d
Two control groups given
distilled water for 12 or
40 d. Two 1% Pb acetate
groups exposed for 12 or
40 d
Day 12: 193.3 pg/L
(19.33 pg/dl)
Day 40: 245.9 pg/L
(24.59 pg/L)
Cardiovascular
histology measured at
12 and 40 d
1% PB acetate group
40 d, M/F, n = 15
BLL = blood lead level; d = day(s); F = female; M = male; Pb = lead; wk = week(s).
4-140
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Table 4-15
Epidemiologic studies of Pb exposure and cerebrovascular disease
Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Menke et al. (2006) NHANES III
n = 13,946
United States
Average individual
NHANES III 1988- born-1946
1994, mortality
follow-up in 2001
Cohort
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
Mean: 2.58
Tertiles:
T1
T2
T3
<1.93
1.94-3.62
>3.63
Age at measurement
Mean: 44.4 yr
Stroke mortality Cox proportional hazard regression
analysis adjusted age, race/ethnicity, sex,
urban residence, cigarette smoking,
alcohol consumption, education, physical
activity, household income, menopausal
status, BMI, CRP, total cholesterol,
diabetes mellitus, hypertension, GFR
category
HR: 1.15 (1.02, 1.28)
Khaliletal. (2009)
Baltimore, MD and
Monongahela Valley,
PA
Study of Osteoporotic
Fractures
n = 533
women, ages 65-
87 yr
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
Mean (SD): 5.3 (2.3)
Range: 1-21
Stroke mortality Cox proportional hazards regression
analysis adjusted forage, clinic, BMI,
education, smoking, alcohol intake,
estrogen use, hypertension, total hip bone
mineral density, walking for exercise, and
diabetes
HR (95% CI) (>8 |jg/dL
blood Pb)b:
1.13 (0.34, 3.81)
Blood Pb measured
1990-1991, mortality
follow-up for -12 yr
Average individual
born -1921
Age at measurement
(Mean): 70
Cohort
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Reference and
Study Design
Study Population Exposure Assessment Outcome
Confounders
Effect Estimates and
95% Clsa
Mousavi-Mirzaei et n = 88
Blood (GFAAS) (pg/dL)
Median (IQR):
6.38 (1.75-34.87)
Acute ischemic
stroke
Logistic regression controlling for lipid
profile and fasting blood sugar
OR (95% CI):
1.04 (1.02, 1.07)
al. (2020)
Birjand, Iran
(44 cases, 44 controls
matched on age and
sex, occupation,
Age at measurement
Mean (SD): 71.95
(11.37)
2016-2017
opium addiction, and
sampling time)
Case-control
Average individual
born -1944
BMI = body mass index; CI = confidence interval; CRP = C-reactive protein; GFAAS = graphite furnace atomic absorption spectrometry; GFR = glomerular filtration rate; HR = hazard
ratio; IQR = interquartile range; NHANES = National Health and Nutrition Examination Survey; OR = odds ratio; Pb = lead; SD = standard deviation; T# = fertile #; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bUnable to be standardized.
4-142
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Table 4-16
Epidemiologic studies of Pb exposure and cardiovascular mortality
Reference and Study
Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Lustberq and Silberqeld
(2002)
United States
NHANES II 1976-1980,
mortality follow-up in
1992
NHANES II
n = 4,190, aged 30-
74
Average individual
born -1924
Blood (GFAAS with
Zeeman correction)15
(Hg/dL)
Mean (SD): 14.0 (5.1)
Median: 13
Tertiles
Circulatory Cox proportional hazard
mortality regression analysis adjusted for
age, sex, location, education,
race, income, smoking, BMI,
exercise
HRC
T2 vs. T1 1.27 (0.97,
T3 vs. T1 1.74 (1.25,
1.57)
2.40)
T1
T2
T3
<10
10-19
20-29
Cohort
Age at measurement:
Mean (SD) 54.1 (13.2)
Schober et al. (2006)
United States
NHANES III 1988-
1994, mortality follow-
up in 2006
-8.5 yr of follow-up
NHANES III
n = 9,686, >40 yr
Average individual
born in or before
-1951
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
T1
T2
T3
<5 (median 2.6)
5-9 (median 6.3)
>10 (median 11.8)
CVD mortality Cox proportional hazard
regression analysis adjusted for
sex, age, race/ethnicity,
smoking, education level.
Did not evaluate BMI nor
comorbidities
HR (95% Cl)c:
CVD
T3 vs. T1 1.55 (1.16, 2.07)
Age at measurement:
>40 yr
Cohort
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Reference and Study
Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Menke et al. (2006)
United States
NHANES III 1988-
1994, mortality follow-
up in 2001
-12 yr of follow-up
Cohort
NHANES III
n = 13,946, >20 yr.
Average individual
born -1946
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
Mean: 2.58
Tertiles:
T1
T2
T3
<1.93
1.94-3.62
>3.63
Age at measurement
Mean: 44.4 yr
CVD, Ml, and
stroke mortality
Cox proportional hazard
regression analysis adjusted
age, race/ethnicity, sex, urban
residence, cigarette smoking,
alcohol consumption,
education, physical activity,
household income,
menopausal status, BMI, CRP,
total cholesterol, diabetes
mellitus, hypertension, GFR
category
HR (95% CI):
CVD: 1.13 (1.06, 1.22)
Ml: 1.19 (1.05, 1.34)
Stroke: 1.15 (1.02, 1.28)
Lanphear et al. (2018)
United States
1988-1994 mortality
follow-up in 2011
-19 yr of follow-up (IQR
17.6-21.0 yr)
Cohort
NHANES III
n = 14,289 >20 yr.
Average individual
born -1947
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
GM: 2.71
Geometric SE: 1.31
10th percentile: 1.0
90th percentile: 6.7
Age at measurement:
Mean: 44.1 yr
CVD, and IHD
mortality
Cox proportional hazards
regression analysis adjusting
forage, sex, household
income, ethnic origin, BMI,
smoking status, alcohol
consumption, physical activity,
concentration of Cd in urine,
BP, healthy eating index
tertiles, HbA1c, and serum
cholesterol
HR (95% CI)
CVD: 1.10 (1.05, 1.15)
IHD: 1.14 (1.08, 1.20)
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Reference and Study
Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
van Bemmel et al.
(2011)
United States
1988-1994, follow-up
through 2007
-7 yr of follow-up for
those with low blood Pb
~7 yr of follow-up for
those with high blood
Pb
NHANES III
n = 3,349 adult age
>40 yr
Average individual
born -1932
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
Median:
<5 |jg/dL: 2.6
>5 |jg/dL: 7.5
Age at measurement
Mean
<5 |jg/dL: 57
>5 |jg/dL: 61
CVD mortality
Cox proportional hazards
adjusting forage, education,
sex, smoking status, and
race/ethnicity
HR (95% CI):
CVD
All: 1.02 (0.93, 1.13)
ALADGG: 1.01 (0.92, 1.10)
ALADCG/GG: 1.13 (0.93, 1.37)
Cohort
Cook et al. (2022)
United States
Baseline 1988-1994,
mortality follow-up
through 2010 (2 yrof
follow-up)
Cohort
NHANES III
n = 15,036
adults >19 y
Average individual
born in or before
-1972
Blood Pb (GFAAS)
(Hg/dL)
Quartiles
Men:
Q1: <2.63
Q2 and Q3:
Q4: >6.23
Women:
Q1: <1.38
Q2 and Q3:
Q4: >3.74
2.63-6.23
1.38-3.74
Age at measurement:
>1 yr
Heart disease
mortality (ICD-10
100-113 I20-I22,
I24, I25-I28,
125.1-125.9, 130-
131, I33, I34-I38,
I40, 142-151, 170-
I78), Ml mortality
(121-122), and
CVD mortality
(heart disease
mortality plus
160-169
(cerebrovascular
disease)
Multivariate Cox model
adjusted for age, gender,
race/ethnicity, family income,
alcohol drinking, cigarette
smoking, BMI, physical activity,
and self-reported health status
at baseline
HR (95% Cl)c:
CVD mortality
Q2 & Q3 vs. Q1: 1.10 (0.84, 1.43)
Q4 vs. Q1: 1.35 (1.03, 1.77)
Per 1 |jg/dL increase in log-
transformed Pb: 1.08 (1.00, 1.16)
Heart disease mortality
Q2 &Q3 vs. Q1: 1.37 (1.04, 1.81)
Q4 vs. Q1: 1.60 (1.21, 2.13)
Per 1 |jg/dL increase in log-
transformed Pb: 1.09 (1.02, 1.16)
Ml mortality
Q2 & Q3 vs. Q1: 1.73 (1.08, 2.79)
Q4 vs. Q1: 1.45 (0.90, 2.32)
Per 1 |jg/dL increase in log-
transformed Pb: 0.95 (0.84, 1.08)
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Reference and Study Exposure
Design Study Population Assessment Outcome Confounders Effect Estimates and 95% Clsa
Ruiz-Hernandez et al.
(2017)
United States
Baseline 1988-1994
and 1999-2004,
mortality follow-up
through 1996 and 2006
Cohort
NHANES III and
NHANES
n = 15,421
age >40yr
Average individual
born in or before
-1951
Blood (|jg/dL)
1988-1994 (GFAAS
with Zeeman
correction)
Mean: 3.2
1999-2004 (ICP-MS)
Mean: 1.9
Age at measurement:
>40 yr
CVD and CHD Poisson regression adjusting
mortality for age, sex, race, smoking
status, physical inactivity,
obesity, hypertension, diabetes,
high total cholesterol, low HDL
cholesterol, lipid lowering
medication, survey period, and
log-transformed urinary Cd
concentrations
RR (95% CI) for twofold increase in
blood Pbc
CVD:
CHD:
1.19 (1.07,
1.24 (1.10,
1.31)
1.41)
(Duan et al.. 2020)
United States
1999-2014, follow-up
through end of 2015
~7 yr of follow-up
NHANES
n = 18,602
Age >20 yr
Average individual
born -1960
Blood (ICP-MS)
(pg/dL)d
Median (IQR)
1.49 (0.93, 2.31)
Age at measurement:
Mean (SD): 45.9 (0.3)
CVD mortality Poisson regression analyses
adjusted for: sex, age, ethnicity,
education, PIR, cotinine
category, BMI, physical activity,
hypertension, and diabetes
RR (95% CI)
CVD: 1.35 (1.15, 1.59)
Cohort
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Reference and Study Exposure
Design Study Population Assessment Outcome Confounders Effect Estimates and 95% Clsa
Aoki etal. (2016)
United States
NHANES
n = 18,602
age >40 yr
Blood (ICP-MS)
(Hg/dL)
Mean (SE): 1.73 (0.02)
1999-2010, follow-up Average individual Age a* measurement
through end of 2011 born-1947 Mean (SE): 57.5 (0.2)
~6 yr of follow-up
Cohort
CVD mortality Cox proportional hazards
(using age during follow-up as
the time scale) adjusting for
race, Hispanic origin, sex,
alcohol consumption, blood Cd,
serum iron, C-Reactive Protein
(CRP), and serum calcium
RR (95% CI) for 10-fold increase in
blood Pbc
Overall: 1.26 (0.91, 1.78)
Control for hematocrit:
1.35 (0.98, 1.86)
Control for hemoglobin:
1.35 (0.98, 1.87)
Hematocrit-corrected:
1.44 (1.05, 1.98)
Hemoglobin-corrected:
1.46 (1.06, 2.01)
Obenq-Gvasi et al.
(2021)
United States
Baseline 1999-2008,
mortality follow-up
through 2014
NHANES
n = 28,852
adults >20 yr
Average individual
born in or before
-1983
Blood Pb (ICP-MS)
(Hg/dL)
Median: 1.55
Age at measurement:
>20 yr
CVD mortality Multivariate Cox model
adjusting for sex, BMI,
smoking, alcohol consumption,
country of birth, and income
*No adjustment for age
HR (>1.55 |jg/dL Blood Pb)c:
2.35 (1.77, 2.93)
Cohort
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Reference and Study
Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Lin etal. (2011)
Taiwan
Years not reported
Cohort (18 mo of follow-
up)
n = 927
Taiwanese adult
patients with end-
stage renal disease
on hemodialysis for
>6 mo, age >18
Baseline blood Pb
(ETAAS) (pg/dL)
Mean: 11.5
Median: 10.4
T1
T2
T3
CVD mortality
<8.51 pg/dL
8.51-12.64 pg/dL
>12.64 pg/dL
Age at measurement
Mean (SD): 55.2 (13.5)
Multivariate Cox model
adjusting forage, previous
cardiovascular diseases
(stroke, Ml, PVD, CHF),
education level, hemodialysis
vintage, using fistula,
normalized protein catabolic
rate, hemoglobin, serum
albumin, creatinine,
cardiothoracic ratio, and
logarithmic transformation of
high-sensitivity CRP
HR (95% CI) (T1: Referent)0
CVD
T2: 3.70 (2.06, 6.48)
T3: 9.71 (2.11, 23.26)
Hemoglobin-corrected:
CVD
T2: 3.52 (0.51, 6.33)
T3: 7.35 (1.64, 33.33)
Weisskopf et al. (2009)
United States
Baseline biomarkers
collected an average of
8 yr prior to death
Cohort
NAS
n = 835
Mostly white elderly
men
Average individual
born in or before
-1940
Blood Pb (GFAAS)
(pg/dL)
Mean (SD): 5.6 (3.4)
Bone Pb (KXRF) (pg/g)
Patella Pb
Mean (SD): 31.2 (19.4)
Tertiles:
T1: <20
T2: 20-31
T3: >31
Tibia Pb
Mean (SD): 21.8 (13.6)
Age at measurement:
Mean (SD): 67 (7)
(13.5)
CVD and IHD Multivariate Cox models. Model
mortality 1 was adjusted for age at
biomarker measurement,
smoking, and education. Model
2 included covariates from
Model 1 as well as mother and
father's occupation, mother and
father's education, occupation
at NAS entry, and salary at
NAS entry. Model 3 includes all
of Model 2 covariates but is
restricted to men < 45 at study
inception. Model 4 is the same
as Model 3, but includes
inverse probability of attrition
rates, as described in the study
HR (95% CI) (3rd vs. 1st fertile)0
Patella Pb:
All men in study (n = 835)
Model 1:
CVD: 1.46 (0.86, 2.48)
IHD: 2.01 (0.86, 4.68)
Model 2:
CVD: 1.45 (0.83, 2.53)
IHD: 2.11 (0.87, 5.13)
Men < 45 at study inception (n = 637)
Model 3:
CVD: 2.23 (1.02, 4.84)
IHD: 4.60 (1.26, 16.8)
Model 4:
CVD: 2.47 (1.23, 4.96)
IHD: 3.09 (1.61, 16.8)
No associations reported between
blood Pb and tibia Pb concentrations
with mortality
4-148
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Reference and Study
Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95% Clsa
Khaliletal. (2009)
Baltimore, MD and
Monongahela Valley,
PA
Blood Pb measured
1990-1991, mortality
follow-up for -12 yr
Study of
Osteoporotic
Fractures
n = 533
Women, ages 65-
87 yr
Average individual
born -1921
Blood (GFAAS with
Zeeman correction)
(pg/dL)
Mean (SD): 5.3 (2.3)
Range: 1-21
Age at measurement
(Mean): 70
CVD, and CHD
mortality
Cox proportional hazards
regression analysis adjusted for
age, clinic, BMI, education,
smoking, alcohol intake,
estrogen use, hypertension,
total hip bone mineral density,
walking for exercise, and
diabetes
HR (95% CI) (>8 pg/dL blood Pb)c
CVD: 1.78 (0.92, 3.45)
CHD: 3.08 (1.23, 7.70)
Cohort
Hollinqsworth and
Rudik (2021) United
States
1999-2016
Quasi-experimental
design
Elderly population
(>65 yr)
Assessed the
change in deaths
(National Vital
Statistics System)
occurring among
this age group
before and after the
phaseout of leaded
gasoline in
professional racing
(NASCAR, ARCA).
County-level blood Pb
measurements in
children
CVD and IHD Difference-in-difference
mortality approach controlling for SES at
the county level (median
income, unemployment rates,
percent minority population),
TRI Pb emissions data
Decline in age-standardized mortality
rate per 100,000 population
CVD:
Race counties: 37
Border counties: 12
IHD:
Race counties: 53
Border counties: 20
Compared mortality
rates in race
counties to
bordering counties
Average individual
born in or before
-1942
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Design Study Population Assessment Outcome Confounders Effect Estimates and 95% Clsa
ARCA = Automobile Racing Club of America; BMI = body mass index; BP = blood pressure; CHD = coronary heart disease; CHF = congestive heart failure; Cd = cadmium;
CI = confidence interval; CRP = C-reactive protein; CVD = cardiovascular disease; ETAAS = electrothermal atomic absorption spectrometry; GFAAS = graphite furnace atomic
absorption spectrometry; GFR = glomerular filtration rate; GM = geometric mean; HDL = high-density lipoprotein; HR = hazard ratio; ICD = International Classification of Diseases;
ICP-MS = inductively coupled plasma mass spectrometry; IHD = ischemic heart disease; IQR = interquartile range; K-XRF = K-shell X-ray fluorescence; Ml = myocardial infarction;
mo = month(s); NAS = Normative Aging Study; NASCAR = National Association for Stock Car Auto Racing; NHANES = National Health and Nutrition Examination Survey; Pb = lead;
PIR = poverty-income ratio; PVD = peripheral vascular disease; Q = quartile; RR = relative risk; SD = standard deviation; SE = standard error; T# = fertile #; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect estimates
are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval.
Categorical effect estimates are not standardized.
bPb analysis method assumed based on data source, not reported in paper.
°Unable to be standardized.
dUnits assumed to be |jg/dL (written as |jg/L in the paper).
4-150
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4.14 References
Aim. J: Kim. NS: Lee. BK: Park. J: Kim. Y. (2018). Association of blood pressure with blood lead and cadmium
levels in Korean adolescents: Analysis of data from the 2010-2016 Korean National Health and Nutrition
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EPA/600/R-23/375
APDA Environmental Protection Januaiy 2024
M m Agency www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 5: Renal Effects
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
-------
CONTENTS
DOCUMENT GUIDE 5-iii
LIST OF TABLES 5-v
LIST OF FIGURES 5-vi
ACRONYMS AND ABBREVIATIONS 5-vii
APPENDIX 5 RENAL EFFECTS 5-1
5.1 Introduction and Summary of the 2013 Integrated Science Assessment 5-1
5.2 Scope 5-3
5.3 Renal Disease and Histology 5-4
5.3.1 Epidemiologic Studies of Kidney Disease 5-5
5.3.2 Toxicological Studies of Kidney Histology 5-10
5.4 Glomerular Filtration Rate and Other Markers of Kidney Function 5-12
5.4.1 Glomerular Filtration Rate 5-13
5.4.2 Albumin, Creatinine, and Albumin-to-Creatinine Ratio 5-17
5.4.3 Uric Acid and Urea 5-21
5.4.4 Proteinuria and Hematuria 5-24
5.4.5 N-Acetyl-p-D-Glucosaminidase and p2-Microglobulin 5-25
5.4.6 Toxicological Studies of Other Indicators of Kidney Function 5-26
5.5 Toxicological Studies of Metal Co-Exposures with Pb 5-27
5.6 Activation of Renin-Angiotensin-Aldosterone System 5-27
5.7 Renal Outcomes Among Children 5-28
5.7.1 Summary of Renal Outcomes Among Children 5-30
5.8 Reverse Causality 5-31
5.8.1 Summary of Reverse Causality 5-33
5.9 Biological Plausibility 5-34
5.10 Summary and Causality Determination 5-37
5.11 Evidence Inventories - Data Tables to Summarize Study Details 5-43
5.12 References 5-77
5-iv
-------
LIST OF TABLES
Table
5-1
Summary of evidence indicating a causal relationship between Pb exposure and renal
effects
5-41
Table
5-2
Epidemioloqic studies of Pb exposure and kidnev disease
5-43
Table
5-3
Animal toxicoloqical studies of Pb exposure and kidnev histoloqv
5-48
Table
5-4
Epidemioloqic studies of Pb exposure and estimated qlomerular filtration rate
5-52
Table
5-5
Animal toxicoloqical studies of Pb exposure and qlomerular filtration rate
5-58
Table
5-6
Epidemiologic studies of Pb exposure and albumin, creatinine, and albumin-to-creatinine
ratio
5-59
Table
5-7
Animal toxicoloqical studies of Pb exposure and albumin and creatinine
5-63
Table
5-8
Epidemioloqic studies of Pb exposure and uric acid3
5-65
Table
5-9
Animal toxicoloqical studies of Pb exposure and measures of uric acid and urea
5-67
Table
5-10
Epidemioloqic studies of Pb exposure and proteinuria and hematuria
5-69
Table
5-11
Epidemioloqic studies of Pb exposure and renal tubular impairment markers3
5-70
Table
5-12
Animal toxicoloqical studies of Pb exposure and other markers of kidnev function
5-71
Table
5-13
Epidemioloqic studies of Pb exposure and renal outcomes in children
5-74
5-v
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LIST OF FIGURES
Figure 5-1 Effect measure modification of association between blood Pb (quartile 1-3 versus quartile
4) and chronic kidney disease incidence. 5-6
Figure 5-2 Kaplan-Meier curve comparing low to high body Pb burden and the development of either
a two-fold increase in serum creatinine from baseline, the need for long-term
hemodialysis, or death among persons with type 2 diabetes. 5-8
Figure 5-3 Association between blood Pb and renal outcomes among patients with type 2 diabetes. 5-9
Figure 5-4 Associations between biomarkers of Pb exposure and estimated glomerular filtration rate. 5-14
Figure 5-5 Association between blood Pb and hyperuricemia among men and women, Korea
National Health and Nutrition Examination Survey, 2016. 5-22
Figure 5-6 Associations between natural log blood Pb (0-4 |jg/dL) and serum uric acid and elevated
serum uric acid (>5.5 mg/g). 5-29
Figure 5-7 Effect measure modification between blood Pb and serum uric acid among adolescents,
National Health and Nutrition Examination Survey 1999-2006. 5-30
Figure 5-8 Locally weighted smoothing plot of adjusted associations between blood Pb levels (with
[left panel] and without [right paned] logarithmic transformation) and serum creatinine. 5-32
Figure 5-9 Potential biological pathways for renal effects following Pb exposure. 5-35
5-vi
-------
ACRONYMS AND ABBREVIATIONS
P2-MG p2-microglobulin
AAS angiotensin-aldosterone system
ACE angiotensin-converting enzyme
ACR albumin-to-creatinine ratio
ALB albumin
AQCD Air Quality Criteria Document
BLB body lead burden
BLL blood lead level
BMI body mass index
BUN blood urea nitrogen
CI confidence interval
CKD chronic kidney disease
CKD-EPI Chronic Kidney Disease Epidemiology
Collaboration
CKiD Chronic Kidney Disease in Children
d day(s)
DKD diabetic kidney disease
EAF electric arc furnace
EBE early biological effect
EDTA ethylenediaminetetraacetic acid
eGFR estimated glomerular filtration rate
ESRD end-stage renal disease
ETAAS Electrothermal Atomic Absorption
Spectrometry
EWAS environment wide association study
F female
FDR false discovery rate
GFAAS graphite furnace atomic absorption
spectrometry
GFR glomerular filtration rate
GW gestational week
ElbAlc hemoglobin Ale
E1DL high-density lipoprotein
hr hour(s)
FIR hazard ratio
ICP-MS inductively coupled plasma mass
spectrometry
IQR interquartile range
ISA Integrated Science Assessment
KIM-1 Kidney Injury Molecule 1
KNHANES Korea National Health and Nutrition
Examination Survey
KRIEFS Korean Research Project on the
Integrated Exposure Assessment to
Hazardous Materials for Food Safety
MCDS Malmo Cancer and Diet Study
MCDS-CC cardiovascular cohort of the Malmo
Cancer and Diet Study
MDRD Modification of Diet in Kidney Disease
mo month(s)
M male
M/F male/female
MONICA Monitory of Trends and Cardiovascular
Disease
NAG N-acetyl-P-D-glucosaminidase
NAS Normative Aging Study
NGAL neutrophil gelatinase-associated
lipocalin
NHANES National Health and Nutrition
Examination Survey
NO3 nitrate
OR odds ratio
Pb lead
PbO lead oxide
Pb(N03)2 lead nitrate
PECOS Population, Exposure, Comparison,
Outcome, and Study Design
PND postnatal day
Q quartile
RAAS renin-angiotensin-aldosterone system
SD standard deviation
SE standard error
SES socioeconomic status
SPHERL Study for Promotion of Health in
Recycling Lead
SUA serum uric acid
T# tertile #
UA uric acid
wk week(s)
WHO World Health Organization
yr year(s)
5-vii
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APPENDIX 5 RENAL EFFECTS
Summary of Causality Determinations for Pb Exposure and
Renal Effects
This appendix characterizes the scientific evidence that supports causality
determinations for lead (Pb) exposure and renal effects. The types of studies evaluated
within this appendix are consistent with the overall scope of the ISA as detailed in the
Process Appendix (see Section 12.4). In assessing the overall evidence, the strengths
and limitations of individual studies were evaluated based on scientific considerations
detailed in Table 12-5 of the Process Appendix (Section 12.6.1). More details on the
causal framework used to reach these conclusions are included in the Preamble to the
ISA (U.S. EPA, 2015). The evidence presented throughout this appendix supports the
following causality conclusion:
Outcome
Causality Determination
Renal Effects
Causal
The Executive Summary, Integrated Synthesis, and all other appendices of this Pb ISA
can be found at https://assessments.epa.gov/isa/document/&deid=359536.
5.1 Introduction and Summary of the 2013 Integrated Science
Assessment
In the 2013 Integrated Science Assessment for Lead (hereinafter referred to as the 2013 Pb IS A;
U.S. EPA, 2013) the epidemiologic and toxicological evidence was judged to be "suggestive of a causal
relationship" between Pb exposures and reduced kidney function among adults. Prospective
epidemiologic studies in adult men in the general population (Tsaih et al„ 2004; Kim et al„ 1996)
supported the temporal relationship between Pb exposure and reduced kidney function at blood lead
levels (BLLs) <10 (ig/dL. As indicated by the male cohort of the Normative Aging Study (NAS), Kim et
al. (1996) noted an increase in serum creatinine with increasing BLLs. Similarly, Tsaih et al. (2004)
indicated a 0.009 mg/dL (95% confidence interval [CI]: -0.0008, 0.0188) annual increase in serum
creatinine over 10 years, with a one-unit increase in natural log tibia Pb. Similar findings were observed
when considering patella Pb as well. These population-based prospective cohort studies showed a
longitudinal association between BLLs and increases in serum creatinine after adjustment for key
potential confounders. In an additional prospective study, higher baseline BLLs were associated with
greater chronic kidney disease (CKD) progression overtime (i.e., reduced estimated glomerular filtration
rate [eGFR] -0.040 mL/min/1.73 m2 [95% CI: -0.0072, -0.008]) in CKD patients (Yu et al.. 2004). Re-
examination of a study from the 2006 Pb Air Quality Criteria Document (AQCD) (U.S. EPA, 2006)
5-1
-------
provided data to conclude that in a population with likely higher past exposures to Pb, a 10-fold increase
in concurrent blood Pb was associated with a decrease in estimated creatinine clearance and that a
3.5 (ig/dL increase in blood Pb had the same negative impact on eGFR as did an increase of 4.7 years in
age or 7 kg/m2 in body mass index (Akesson et al.. 2005). Cross-sectional studies of the general adult
population added support to the associations observed in prospective epidemiologic studies. The majority
of cross-sectional studies reported associations between higher measures of Pb exposure and impaired
renal function (Navas-Acicn et al.. 2009; Muntner et al.. 2005; Muntner et al.. 2003). Other studies in
clinical trials of CKD patients treated with ethylenediaminetetraacetic acid (EDTA) chelation provide
supportive results; however, these studies had uncertainties concerning small sample sizes and lack of
researcher blinding.
With respect to the animal toxicology evidence, the 2013 Pb ISA noted that at BLLs >30 (ig/dL,
there was clear evidence that Pb exposure caused changes to kidney morphology and function (Khali 1-
Manesh et al.. 1992b; Khalil-Manesh et al.. 1992a). Evidence for functional changes in animals following
lower Pb exposures resulting in BLLs <20 (ig/dL was generally not available. At BLLs between 20 and
30 (ig/dL, studies with various exposure scenarios and in various lifestages provided evidence for reduced
kidney function measures (e.g. decreased creatinine clearance, increased serum creatinine, increased
blood urea nitrogen [BUN]). In addition, previous reviews have clearly established that exposure to Pb
can result in the production of reactive oxygen species and markers of inflammation in the blood or
kidneys over a similar range of BLLs (see (U.S. EPA. 2013)).
However, there were important uncertainties identified in the 2013 Pb ISA. First, because
epidemiologic studies report effects in adult populations with past Pb exposures that are likely higher,
uncertainty exists as to the Pb exposure level, timing, frequency, and duration contributing to the
associations observed with blood or bone Pb levels. Second, due to the kidney's role in removing toxins
from the blood, it is plausible that reverse causality could explain the associations observed in
epidemiologic studies. While the epidemiologic and animal toxicological studies mentioned above
suggest that reverse causality does not contribute substantially to associations between higher BLLs and
reduced kidney function, reverse causation remained a plausible hypothesis. Thus, this bidirectional
relationship is possible and additional evidence was needed to fully elucidate the extent to which
diminished kidney function may itself result in increased blood or bone Pb levels.
When considered as a whole, although there was evidence of impaired kidney function in some
epidemiologic studies, as well as animal toxicological evidence of oxidative stress and impaired kidney
function providing biological plausibility for those associations, important uncertainties remained. In
particular, uncertainties related to the potential for reverse causality in epidemiologic studies and the lack
of animal toxicological studies indicating impaired kidney function at lower BLLs were noted. As a
result, the relationship between Pb exposure and reduced kidney function was judged to be suggestive of
a causal relationship.
5-2
-------
The following sections provide an overview of study inclusion criteria for this Appendix
(Section 5.2), an evaluation of the health evidence published since the 2013 Pb ISA (Sections 5.3-5.8), a
summary of the biologically plausible pathways by which exposure to Pb could result in the health
outcomes observed in epidemiologic studies (Section 5.9), a discussion of the causal determination for Pb
exposure and renal effects (Section 5.10), tables providing toxicological and epidemiologic study-specific
details (Section 5.11), and references (Section 5.12).
5.2 Scope
The scope of this appendix is defined by Population, Exposure, Comparison, Outcome, and Study
Design (PECOS) statements. The PECOS statements define the objectives of the review and establish
study inclusion criteria, thereby facilitating identification of the most relevant literature to inform the Pb
ISA.1 In order to identify the most relevant literature, the body of evidence from the 2013 Pb ISA was
considered in the development of the PECOS statements for this Appendix. Specifically, well-established
areas of research; gaps in the literature; and inherent uncertainties in specific populations, exposure
metrics, comparison groups, and study designs identified in the 2013 Pb ISA inform the scope of this
Appendix. The 2013 Pb ISA used different inclusion criteria than the current ISA, and the studies
referenced therein often do not meet the current PECOS criteria (e.g. due to higher or unreported
biomarker levels). Studies that were included in the 2013 Pb ISA, including many that do not meet the
current PECOS criteria, are discussed in this appendix to establish the state of the evidence prior to this
assessment. With the exception of supporting evidence used to examine the biological plausibility of Pb-
associated renal effects, recent studies were only included if they satisfied all of components of the
following discipline-specific PECOS statements:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect;
Exposure: Exposure to Pb2 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
'The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g. involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g. that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
2Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area that was of
particular relevance to the NAAQS review (e.g. longitudinal studies designed to examine recent versus historical Pb
exposure).
5-3
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exposure3; or intervention groups in randomized trials and quasi-experimental studies;
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g. per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles);
Outcome: Renal effects including, but not limited to, renal function and CKD; and
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g. mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages);
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a BLL of 30 (ig/dL or below;4,5
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control;
Outcomes: Renal effects; and
Study Design: Controlled exposure studies of animals in vivo.
5.3 Renal Disease and Histology
The primary function of the kidneys is to filter waste from the body while maintaining
appropriate levels of water and essential chemicals, such as electrolytes. Kidney disease occurs when
kidney function becomes impaired and cannot perform these functions adequately. Section 5.3.1 evaluates
the epidemiologic evidence for kidney disease and exposure to Pb. Kidney disease is often accompanied
by changes in the structure of the kidney. For example, glomerular or tubular hypertrophy can be used as
an indication of kidney dysfunction and disease. Similarly, changes in the number or morphology of renal
tubules or podocytes can also be indicative of kidney disease. Thus, Section 5.3.2 presents the animal
3Studies that estimate Pb exposure by measuring Pb concentrations in PMio and PM2.5 ambient air samples are only
considered for inclusion if they also include a relevant biomarker of exposure. Given that size distribution data for
Pb-PM are fairly limited, it is difficult to assess the representativeness of these concentrations to population
exposure [Section 2.5.3 (U.S. EPA. 2013)1. Moreover, data illustrating the relationships of Pb-PMio and Pb-PNL 5
with BLLs are lacking.
4Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
5This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLL.
The 95th percentile of the 2011-2016 National Health and Nutrition Examination Survey distribution of BLL in
children (1-5 years; n = 2,321) is 2.66 (ig/dL (Eganet al„ 2021) and the proportion of individuals with BLL that
exceed this concentration varies depending on factors including (but not limited to) housing age, geographic region,
and a child's age, sex, and nutritional status.
5-4
-------
toxicological studies that have examined histological sections of kidneys for abnormalities and changes in
structure following Pb exposure.
5.3.1 Epidemiologic Studies of Kidney Disease
The 2013 Pb ISA (U.S. EPA, 2013) and 2006 Pb AQCD (U.S. EPA, 2006) highlighted several
studies indicating an association between biomarkers of Pb exposure and indicators of decreased renal
function and progression of CKD. Several recent studies specifically evaluated biomarkers of reduced
kidney function and the development of CKD or end-stage renal disease (ESRD). Study-specific details,
including Pb biomarker levels, study population characteristics, potential confounders, and select results
from these studies are highlighted in Table 5-2. Study details in Table 5-2 include standardized results
(kidney disease associated with a 1 (ig/dL increase in BLL or a 10 |ig/g increase in bone Pb level) as well
as results that could not be standardized with the information provided in each paper.
5.3.1.1 Chronic Kidney Disease
The 2013 Pb ISA presented a number of occupational studies evaluating the association between
Pb exposure and CKD, but the results were relatively inconsistent. More recent evidence helps to
disentangle the evidence previously presented and consistently indicates an association between
biomarkers of Pb exposure and CKD development. A study among the cardiovascular cohort of the
Malmo Cancer and Diet Study (MCDS-CC) in Malmo, Sweden evaluated the development of CKD by
assessing baseline BLLs (obtained in 1991-1994) and incident CKD (assessed in 2007-2012) (Harariet
al„ 2018). In this study, CKD was confirmed through medical records. When each individual quartile of
blood Pb was compared with the lowest, there was no association with incident CKD. However, when the
three lower quartiles (Q1-Q3 median 2.2 (ig/dL) were compared with the highest (Q4 median 4.6 |ig/dL).
an association was observed between blood Pb and CKD (hazard ratio [HR]: 1.49 [95% confidence
interval (CI): 1.07, 2.08]), while controlling for baseline eGFR in the models. This association remained
stable even after stratification by several covariates (Figure 5-1).
5-5
-------
Covariates
Overall
Age
Sex
Smoking
Alcohol intake
Waist circumference
Diabetes mellitus
Hypertension
Low Education
Categories
Overall
<58 years
>58 years
Female
Male
Never
Former
Current
<6.6 ^ day
>6.6 g day
<82 cm
>82 cm
No
Yes
No
Yes
No
Yes
Hazard Ratios
(95% CI)
1.49 (1.07,
1.21 (0.64,
1.61 (1.09,
1.41 (0.73,
1.58 (1.06,
1.42(0.78,
1.36(0.82,
1.56 (0.82,
1.46 (0.87,
1.53 (0.97,
1.61 (0.68,
1.46(1.01,
1.62 (1.10,
0.91 (0.44,
0.88 (0.31,
1.57(1.11,
1.79(1.07,
1.29 (0.83,
2.08)
2.28)
2.37)
2.73)
2.35)
2.58)
2.26)
2.98)
2.45)
2.41)
3.83)
2.11)
2.38)
1.89)
2.47)
2.23)
2.99)
2.00)
1
1.5
i—i—i—r
2.5 3 3.5 4
P-value
for
interaction
0.4
0.4
0.9
0.9
0.7
0.2
0.6
0.3
.6 .7 .8 .9 1
Hazard ratios (95% CI)
Cm = centimeters; g = grams.
Source: Harari et al. (2018).
Figure 5-1 Effect measure modification of association between blood Pb
(quartile 1-3 versus quartile 4) and chronic kidney disease
incidence.
A case-control study in Taiwan matched healthy controls by age and sex to those with CKD (Wu
et al.. 2019). In this study, CKD was defined as an eGFR <60 mL/min/1.73 m2 for at least 3 consecutive
months. When compared with the lowest tertile (<2.784 (ig/dL) of blood Pb, the odds of CKD increased
with each increasing tertile of red blood cell Pb. Compared with the lowest tertile, the highest tertile
(>4.635 (ig/dL) of blood Pb, was associated with an odds ratio (OR) of 6.48 (95% CI: 3.23, 12.99) for
CKD. Since Pb can lead to oxidative damage in the kidney, the authors tested the association between red
blood cell Pb and CKD modified by selenium. Selenium has antioxidant properties and selenium
homeostasis is maintained by the kidney. When examined, serum selenium appeared to modify this
association.
A large National Health and Nutrition Examination Survey (NHANES, 1999-2016) analysis
included an environment wide association study (EWAS) on 262 environmental chemicals (Lee et al..
2020). Individual CKD components including albuminuria (urinary albumin [ALB]-to-creatinine ratio
[ACR] >30 mg/g) and reduced eGFR (<60 mL/min/1.73 m2 based on the Chronic Kidney Disease
5-6
-------
Epidemiology Collaboration (CKD-EPI) calculation) and a set of composite CKD measures were used as
outcome measures in this study. A discovery data set was created by combining five NHANES cycles
(1999-2000, 2003-2004, 2007-2008, 2011-2012, and 2015-2016). Individual regression analyses were
conducted for each survey cycle and combined using a random-effects meta-analysis. Chemicals with a
false discovery rate (FDR) <1% in the meta-analysis were considered as potential risk factors for CKD.
Identified chemicals were then reanalyzed in the rest of the survey cycles (2001-2002, 2005-2006, 2009-
2010, and 2013-2014) and referred to as the "validation" set. Blood Pb was analyzed in both the
discovery and validation sets for reduced eGFR and the composite CKD definitions (the FDR for
albuminuria was >1%). When assessing a composite CKD measurement (ACR >300 mg/g and eGFR
<60 mL/min/1.73 m2, ACR >30 mg/g and eGFR <45 mL/min/1,73m2 or eGFR <30 mL/min/1.73 m2),
there was a positive association with blood Pb in the discovery (OR: 1.73 [95% CI: 1.54, 1.95]) and
validation (OR: 1.61 [95% CI: 1.35, 1.901) sets. In contrast. Kim et al. (2015) cross-sectionallv evaluated
the association between blood Pb and self-reported CKD using the Korea National Health and Nutrition
Examination Survey (KNHANES 2011) and indicated a null association (OR: 1.05 [95% CI: 0.85, 1.30])
after controlling for confounders. The association remained null after stratifying by diabetic status.
5.3.1.2 End-Stage Renal Disease
ESRD is diagnosed when CKD progresses to a level in which renal replacement therapy
(hemodialysis or transplantation) is required. Sommar et al. (2013) combined studies including the
Northern Sweden Health and Disease Study and MCDS. The Northern Sweden Health and Disease Study
incorporates data from three different cohorts: the Vasterbotten Intervention Project, the Northern Sweden
World Health Organization (WHO) Monitory of Trends and Cardiovascular Disease (MONICA) study,
and Mammography Screening Project. All included studies collected baseline data on erythrocyte Pb
levels. Cases of ESRD were identified through the Swedish Renal Registry and linked with erythrocyte
Pb data from the above cohorts. Controls were selected from within each of the respective studies and
were matched on age, sex, cohort, and time of sampling. In a combined cohort of over 130,000
individuals, 118 cases of ESRD were identified (with 378 controls). Here, a one-unit (fig/dL) increase in
erythrocyte Pb was associated with an OR of 1.14 (95% CI: 1.03, 1.26).
5.3.1.3 Diabetic Nephropathy
Diabetic nephropathy refers to a reduction in kidney function leading to ESRD among those with
type I or II diabetes mellitus. The development of ESRD or CKD is more likely among those with
diabetes mellitus, compared with the general population. Huang et al. (2013) evaluated body lead burden
(BLB) and blood Pb among persons with type 2 diabetes with stage 3 diabetic nephropathy (eGFR range:
30-60 mL/min/1.73 m2). A combination of X-ray fluorescence detecting bone Pb concentrations and
calcium disodium EDTA demobilization tests are typically used to assess BLB. Typically, a BLB <80 |ig
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is considered to be within the normal range, while a BLB >600 |ig is equivalent to Pb poisoning. This
small study (n = 89) indicated a decrease in eGFR associated with a one-unit increase in either BLB
(-0.022 ml./inin/1.73 m2 [95% CI: -0.039, -0.005]) or blood Pb (-0.298 mL/min/1.73 m2 [95% CI:
-0.525, -0.071]). Additionally, there was an increased risk of the "primary outcome" (either a two-fold
increase in serum creatinine from baseline, the need for long-tenn hemodialysis, or death) with a one-unit
increase in Pb BLB (HR: 1.01 [95% CI: 1.01, 1.02]), and a BLB between 80-600 fig was associated with
an HR of 2.79 (95% CI: 1.25, 6.25). The Kaplan-Meier analysis conducted within this study demonstrated
that diabetic patients with higher BLB (>80 |ig) were more likely to reach the primary outcome at an
accelerated rate compared with those with lower BLB (Figure 5-2).
1 -
0.8 -
"3
0.6 -
J-
I 0.4 -
—>
V
0.2 -
0-
Cum. survival (low-normal BLB) Cum. survival (high-normal BLB)
# Event times (low-normal BLB) ^ Event times (high-normal BLB)
BLB = body lead burden.
Source: Huang et al. (2013).
Figure 5-2 Kaplan-Meier curve comparing low to high body Pb burden and
the development of either a two-fold increase in serum creatinine
from baseline, the need for long-term hemodialysis, or death
among persons with type 2 diabetes.
11111 I ¦ i i i | ¦ ¦ ¦ ' 1 1 1 ' » I 1 1 ' » 1
A a—i 4—.
I
| i i i i | i i i i | i i i i | i i i i | i i i i |
0 5 10 15 20 25
Time (month)
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A recent cross-sectional study evaluated diabetic kidney disease (DKD) in those with type 2
diabetes (Hagedoorn et al.. 2020). The authors directly calculated glomerular filtration rate (GFR) by
measuring creatinine in a 24-hour urine sample, rather than calculating eGFR from a single serum
creatinine measurement. In addition, the study also evaluated albuminuria, defined as a 24-hour urinary
ALB excretion >30 mg/day. Each doubling of blood Pb (on a log2 scale) was associated with an OR of
1.83 (95% CI: 1.07, 3.15) for creatinine clearance <60 mL/min/1.73 m2 and an OR of 1.75 (95% CI: 1.11,
2.74) for albuminuria. Wan et al. (2021) cross-sectionally evaluated diabetic patients in China by
evaluating the association between BLLs and both an ACR >30 mg/g and DKD (defined as ACR
>30 mg/g or eGFR <60 mL/min/1.73 m2). When comparing the highest quartile of blood Pb (>3.7 (ig/dL)
with the lowest quartile of blood Pb (<1.8 (.ig/dL). the odds of an elevated ACR (>30 mg/g) (OR: 1.31
[95% CI: 1.02, 1.69]) and the presence of DKD (OR: 1.36 [95% CI: 1.06, 1.74]) were higher. The dose
response indicated increased odds of DKD with increasing blood Pb and a decrease in eGFR with each
quartile increase in BLL (Figure 5-3).
High ACR
P for trend ¦ 0.026
One In BLL 5D increment
BLL Quartile 4
BLL Quartile 3
ELL Quartile 2
BLL Quartile 1
0.5 1.0 1.5
Odds ratio (95%CI)
2.0
DKD
One In BLL SD increment
BLL Quartile 4
BLL Quartile 3
BLL Quartile 2
BLL Quartile 1
Pfor Irend = 0.011
0.5 1.0 1.5
Odds ratio (95%CI)
2.0
Ln ACR
Pfor trend < 0.001
One In BLL SD increment
BLL Quartile 4
BLL Quartile 3
BLL Quartile 2
BLL Quartile 1
-0,1 0.0 0,1 0.2 0.3 0,4
regression coefficients (95%CI)
eGFR
Pfor trend < 0.001
One In BLL SD increment
BLL Quartile 4
BLL Quartile 3
BLL Quartile 2
BLL Quartile 1
-6.0 -1.0 -2.0 0 2.0
regression coefficients (95%CI)
ACR = aibumin-to-creatinine ratio; BLL = blood lead level; DKD = diabetic kidney disease; eGFR = estimated glomerular filtration
rate; SD = standard deviation.
Source: Wan et al. (2021).
Figure 5-3 Association between blood Pb and renal outcomes among
patients with type 2 diabetes.
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5.3.1.4
Nephrolithiasis
Nephrolithiasis, or kidney stones, can be the result of a disruption in calcium homeostasis.
Exposure to Pb, a nephrotoxicant, can potentially compete with calcium in binding to calcium-binding
receptors, leading to the development of kidney stones. A prospective study evaluated baseline BLLs and
the development of incident nephrolithiasis (verified by medical records) in a Flemish population as part
of the Cadmium in Belgium (CadmiBel) study (Hara et al.. 2016). Baseline blood Pb measurements were
obtained between 1985 and 1989, and the incidence of nephrolithiasis was measured through October
2014. Approximately half of the population (747 out of 1302) had a second blood Pb measurement
between 1991 and 2004. According to the baseline measurement, there was an increased risk of incident
nephrolithiasis for each doubling of blood Pb (HR: 1.35 [95% CI: 1.06, 1.73]). A similar risk was noted
when averaging the baseline and the follow-up BLLs (HR: 1.32 [95% CI: 1.03, 1.71]). Furthermore,
applying an additional regression dilution bias correction increased the magnitude of the baseline
association (HR: 1.44 [95% CI: 1.07, 1.93]). Conversely, an NHANES (2007-2016) analysis cross-
sectionally assessed the association between the self-reported prevalence of kidney stones and BLLs (Sun
et al.. 2019). Compared with the lowest or referent group (blood Pb: 0.05 (.ig/dL). increasing blood Pb
values corresponded to ORs indicative of a protective effect against kidney stones in this population, with
the highest blood Pb group (>5 (ig/dL) corresponding to an OR of 0.64 (95% CI: 0.46, 0.90). This
association persisted even when stratifying by sex, ethnicity, and body mass index (BMI).
5.3.1.5 Summary of Kidney Disease
The sections above present mostly positive associations between BLLs and some kidney diseases
from epidemiologic studies. More specifically, all but a single epidemiologic study demonstrated a
positive association between measures of body Pb and some measure of CKD (Lee et al.. 2020; Wu et al..
2019; Harari et al.. 2018). ESRD (Sommar et al.. 2013). and diabetic nephropathv(Wan et al.. 2021;
Hagedoorn et al.. 2020; Huang et al.. 2013). However, evidence for an association between measures of
Pb and nephrolithiasis (i.e., kidney stones) was limited to a couple of studies with conflicting results (Sun
et al.. 2019; Hara et al.. 2016). Importantly, epidemiologic studies demonstrating positive associations
between measures of Pb and kidney disease were conducted in a variety of geographical areas and in
different study populations. Moreover, in general, these analyses also controlled for a number of potential
confounders, thus increasing confidence in these associations.
5.3.2 Toxicological Studies of Kidney Histology
In previous Pb ISAs, some studies reported that exposure to Pb induced changes in renal
structure. For example, Roncal et al. (2007) found that Pb increased tubulointerstitial injury and
arteriolopathy in rats. The BLL in this study was 26 (ig/dL. Moreover, Jabeen et al. (2010) reported that
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Pb exposure (no specified BLL) decreased kidney cortical thickness, decreased the diameter of renal
corpuscles, and increased renal tubular atrophy in mice. In contrast to these studies, Vvskocil et al. (1995)
reported that Pb exposure to female rats resulting in a BLL of 36 (ig/dL caused no change in kidney
function or nephrotoxicity. More information on these and other studies examining renal effects following
Pb exposure can be found in Table 4-28 of the 2013 Pb ISA (U.S. EPA. 2013).
A number of studies published since the 2013 Pb ISA with BLLs <30 (ig/dL have examined
kidney tissue for indications of abnormalities following Pb exposure by drinking water or gavage. Basgen
and Sobin (2014) reported that in young mice, exposure to Pb leading to BLLs ranging from 2.74 (ig/dL
to 4.7 (ig/dL resulted in a statically significant glomerular volume increase (p <0.05), but similar numbers
of podocytes and podocyte volume densities. At higher BLLs (11.7 (ig/dL to 20.3 (.ig/dL). this change to
kidney structure was not observed. With respect to glomerular components at lower BLLs, the authors
reported a statistically significant effect (p <0.05) on mesangial volume and capillary lumen volume, but
not podocyte volume (Basgen and Sobin. 2014). Similarly, although control rats had a well-preserved
nucleus and normal tubular and glomerular morphology, renal tubules from rats exposed to Pb
(21.9 (ig/dL BLL) had irregular cell shapes, changes in cell and nuclear sizes, and minimal amounts of
cytoplasm (Alcaraz-Contreras et al.. 2016). Cells from renal tubules also displayed a loss of apical
microvilli (Alcaraz-Contreras et al.. 2016). In an additional study, Pb exposure resulting in a BLL of
-12 (ig/dL on postnatal day (PND) 21 and -23 (ig/dL on PND 30 resulted in a statistically significant
decrease (p = 0.01) of 1-a-hydroxylase at PND 21, but not PND 30 by western blot relative to controls.
These authors further noted that the western blot results were in agreement with immunohistochemistry
on kidney cells. (Rahman et al.. 2018). Shi et al. (2020) reported that kidney tissue from rats with a BLL
of -10.21 (ig/dL displayed cellular debris, tubular dilation, glomerulus hypercellularity, and other signs of
distress while control kidney tissue showed no major histopathological changes. In agreement with this
study, Laamech et al. (2016) also reported that relative to control animals, Pb-treated mice with a BLL of
18 (ig/dL displayed glomerular hypercellularity. Gao et al. (2020) similarly reported histopathological
changes consistent with damage following Pb exposure in rats. In this study, the BLL was 10.6 (ig/dL and
the authors reported congestion and vasodilation of the renal interstitium and swelling of tubules in Pb-
exposed animals while controls appeared to have normal kidney structure (Gao et al.. 2020). Likewise, Li
et al. (2017) reported hyperemic glomeruli, increased glomerular volume, and swelling of some renal
tubular epithelial cells after Pb exposure resulting in an average BLL of -30 (ig/dL, while histological
sections from the control mice were normal.
In addition to the drinking water and gavage studies described above, Andielkovic et al. (2019)
reported that Pb exposure (BLL -30 (ig/dL) resulted in acute passive kidney hyperemia, but no significant
pathologic changes following gavage. Moreover, Carlson et al. (2018) reported that in mice, exposure to
Pb by drinking water resulted in minor renal lesions that were similar to those in control mice (e.g. simple
tubular hyperplasia) and that these lesions were not indicative of major systemic health problems.
However, it is worth noting that the BLL in this study was only 2.89 (ig/dL, and thus, extensive renal
lesions may not be expected.
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In addition to the analyses described above, recent studies have examined the effects of Pb after
inhalation exposure. Following inhalation exposure to Pb-oxide nanoparticles for 6 weeks (resulting in
-14 (ig/dL BLL), Dumkova et al. (2017) reported minor changes in kidney appearance relative to some,
but not all control mice. These changes were mainly areas of mild inflammation around the renal
corpuscles and tubules. Similarly, ultrastructural analysis of the kidneys also revealed only minor
differences between Pb-treated and control mice. However, these authors noted thicker lamina densa, and
the average distance between endothelial cell and podocyte cytoplasmic membranes increased following
Pb exposure (Dumkova et al.. 2017). In an additional study by the same author, Dumkova et al. (2020a)
used Pb nitrate nanoparticles and a longer inhalation time (11 weeks, resulting in a BLL of 8.5 (ig/dL) and
reported that there were obvious morphological changes in renal tubules when compared with control
mice. Moreover, pedicles of podocytes were reported to be irregularly arranged or lost altogether. After a
5-week clearance period, Pb levels in the kidneys and blood declined substantially (BLL 1 |ig/dL). and
there was evidence of regeneration in tubular and glomerular kidney tissue. However, in another analysis
by the same author, Dumkova et al. (2020b) reported that an 11-week exposure with Pb-oxide
nanoparticles resulted in no significant change in mouse kidney morphology when compared with
controls with BLLs as high as 17 (ig/dL. Thus, it is possible that exposure to different forms of Pb results
in differing degrees of kidney damage, but additional studies would be needed to confirm this possibility.
When considered as a whole, substantial evidence exists from studies published since the last Pb
ISA suggesting that exposures resulting in BLLs <30 (ig/dL result in histological abnormalities in the
kidneys. Moreover, these abnormalities were reported following all tested routes of Pb exposure
(i.e., drinking water, gavage, and inhalation). Additional information on the experimental design of more
recent renal histology studies can be found in Table 5-9.
5.3.2.1 Summary of Kidney Histology Studies
Most animal toxicological studies demonstrate that exposure to Pb results in abnormalities or
damage to kidney cells or tissue (see Section 5.3.2). Effects include changes in glomerular and nucleus
morphology, as well as changes in the amount of cellular cytoplasm. Histological effects were also
identified following both oral and inhalation exposures.
5.4 Glomerular Filtration Rate and Other Markers of Kidney
Function
The gold standard for assessing kidney function involves measurement of the GFR through
administration of an exogenous radionuclide or radiocontrast marker (e.g. 1251-iothalamate, iohexol)
followed by timed sequential blood samples or, more recently, kidney imaging, to assess clearance
through the kidneys. This procedure is invasive and time-consuming. Therefore, serum levels of
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endogenous compounds are routinely used to estimate GFR in large epidemiologic studies and clinical
settings. Creatinine is the most commonly measured endogenous compound; measures of urea (e.g. BUN)
and uric acid (UA) have also been examined for this purpose. Increased serum concentration or decreased
kidney clearance of these markers can indicate kidney dysfunction. The main limitation of endogenous
compounds identified to date is that non-kidney factors impact their serum levels. Specifically, since
creatinine is derived from creatinine in muscle, muscle mass and diet affect serum levels resulting in
variations in different population subgroups (e.g. women and children compared with men) that are
unrelated to kidney function. Measured creatinine clearance, involving measurement and comparison of
creatinine in both serum and urine, can address this problem. However, measured creatinine clearance
utilizes timed urine collections, traditionally over a 24-hour period, and the challenge of complete urine
collection over an extended time period makes compliance difficult. Therefore, equations to estimate
kidney filtration that utilize serum creatinine but also incorporate age, sex, race, and, in some cases,
weight (in an attempt to adjust for differences in muscle mass) have been developed. Although these are
imperfect surrogates for muscle mass, such equations are currently the preferred outcome assessment
method.
5.4.1 Glomerular Filtration Rate
5.4.1.1 Epidemiologic Studies of Estimated Glomerular Filtration Rate
Glomerular filtration rate can be estimated based on a variety of different biological factors and
measured kidney function markers. An equation from the Modification of Diet in Kidney Disease
(MDRD) Study (Levey et al„ 2000; Levey et al„ 1999) calculates eGFR based on serum creatinine, race,
sex, and age. With widespread use of the MDRD equation, it became clear that the equation possibly
underestimates GFR at high levels, even in the normal range. A second, creatinine-based equation,
CKD-EPI, was recently developed in order to be more accurate than the MDRD equation, particularly at
higher GFRs. This equation also incorporates serum creatinine, race, sex, and age. However, both
equations do not consider an adjustment for muscle mass, therefore alternative biomarkers, such as
cystatin C, a cysteine protease inhibitor that is filtered, reabsorbed, and catabolized in the kidney (Fried,
2009), have also been developed. The normal range of GFR is between 90 and 120 mL/min/1.73 m2, and
GFR <60 mL/min/1.73 m2is typically indicative of kidney disease, while GFR <15 mL/min/1.73 m2 is a
marker of renal failure.
The 2006 Pb AQCD (U.S. EPA, 2006) and the 2013 Pb ISA (U.S. EPA, 2013) considered
several studies that evaluated associations between biomarkers of Pb exposure and eGFR specifically in
healthy adult populations as well as populations with comorbid conditions. Most studies indicated a
relationship between Pb biomarkers and decreases in eGFR. Yu et al. (2004) studied eGFR (MDRD)
among CKD patients in Taiwan and reported an association between blood Pb and an accelerated
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decrease in eGFR. The 2013 Pb ISA (U.S. EPA. 2013) highlighted several cross-sectional studies that
examined the association between BLLs and either eGFR or creatinine clearance. Navas-Acien et al.
(2009) evaluated the association between BLLs and reduced eGFR (eGFR <60 mL/min/1.73 m2)
measured with the MDRD equation. This study reported reduced eGFR for the highest quartiles of blood
Pb (>2.4 (.ig/dL). compared with the lowest (<1.1 (ig/dL). Several recent studies have also longitudinally
and cross-sectionally evaluated the association between blood Pb and various measures of eGFR. Study-
specific details, including BLLs, study population characteristics, confounders, and select results from
these studies are highlighted in Figure 5-4 and Table 5-4. Studies in Figure 5-4 are standardized to be
interpreted as changes in eGFR associated with a 1 (ig/dL increase in BLL. Study details in Table 5-4
include standardized results as well as results that could not be standardized using the information
provided in each paper.
Study Population Pb distribution Period
(Hg/m3)
Yu et al. 2004 Adult CKD patients Mean (SD): 4.2 (2.2) 4 Years •
Chung et al. 2020 Community residents living near an Geometric Mean (IQR) ~5 Years •
a electric arc furnace (EAF) v '
Distance from EAF
< 500 m: 2.41 (1.22-6.19)
500-1000 m: 2.26 (1.16-4.83)
1000-1500 m: 2.12 (1.05-4.67)
1500-2000 m: 2.23 (0.98-4.31)
> 2000 m: 2.03 (1.03-4.31)
Chung et al. 2014 KNHANES Geometric mean: 2.5 Concurrent —•—
-8 -7 -6 -5 -4 -3 -2 -1 0
Change in eGFR (mL/min/1.73 mA2 body surface area) (95% CI)
per 1 ug/dL increase in blood Pb
CKD = chronic kidney disease.
Note: Studies published since the 2013 Pb ISA. Associations presented per 1 |jg/dL increase in BLL
Figure 5-4 Associations between biomarkers of Pb exposure and estimated
glomerular filtration rate.
A recent analysis of the cardiovascular cohort of the Malmo Diet and Cancer Study (MDCS-CC)
evaluated the change in eGFR from baseline to follow-up (Harari et al.. 2018). Study participants were
initially recruited between 1991 and 1994 (mean age: 57), when BLLs and eGFR (CKD-EPI) were
initially assessed. Follow-up occurred between 2007 and 2012 (mean age: 73), when eGFR was re-
assessed. Compared with the lowest quartile of blood Pb (Ql: 0.15-1.85 (.ig/dL). eGFR was reduced in
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each of the higher quartiles. The highest quartile of blood Pb (3.3-25.8 (ig/dL) was associated with a
2.3 mL/min/1.73 m2 decrease (95% CI: -3.8, -0.73) in eGFR. Another recent longitudinal study of eGFR
and BLLs was conducted in China (Liu et al.. 2020). This study, among middle aged and older adults,
evaluated the annual decline in eGFR (CKD-EPI) from baseline (2010) through the final follow-up
(2013). Annual decline in eGFR was calculated as follows: Baseline eGFR - Follow-up eGFR)/Years of
follow-up. Compared with the lowest quartile (<0.843 (ig/dL) there was a 0.83 mL/min/1.73 m2 (95% CI:
0.31, 1.35) decline in eGFR for those in the highest quartile (>1.895 (ig/dL) per year. In addition, Chung
et al. (2020) described a longitudinal cohort of those living near an electric arc furnace (EAF) in Taiwan.
This study evaluated blood Pb assessed at baseline (measured in 2010-2011) and eGFR (method not
specified; measured in 2015-2016). At follow-up, every 1 (ig/dL increase in blood Pb was associated with
a 2.25 mL/min/1.73 m2 (95% CI: -3.50, -1.01) decrease in eGFR. A smaller prospective cohort study
(BioCycle study) evaluated several kidney markers, including eGFR (MDRD) among premenopausal
women in the United States (Pollack et al.. 2015). The BioCycle study followed women for 2 menstrual
cycles and included a total of 16 clinic visits (8 per cycle) timed to certain days of the menstrual cycle.
Each doubling of blood Pb was associated with a -3.73% change in eGFR (95% CI: -6.55, -0.83).
However, there was no association between a doubling of blood Pb and either eGFR <90 mL/min/1.73 m2
(OR: 0.32 [95% CI: 0.08, 1.21]), or <60 mL/min/1.73 m2 (OR: 0.32 [95% CI: 0.08, 1.21]).
Other studies evaluating biomarkers of Pb exposure and renal function were cross-sectional in
nature. Cross-sectional studies can be useful for determining associations but are unable to establish the
temporality of the association. Recently, the Study for Promotion of Health in Recycling Lead
(SPHERL), a cross-sectional study evaluating newly hired Pb workers at battery manufacturing and Pb
recycling plants in the United States, assessed blood Pb (taken at baseline, before large potential
occupational exposure) and concurrent eGFR (CKD-EPI) (Muiai et al.. 2019). However, the SPHERL
study only included men (n = 447) and indicated null associations with eGFR, whether using creatinine,
cysteine C, or a combination of creatinine and cystatin C with increasing BLLs.
In addition, several nationally representative studies (KNHANES, NHANES) also evaluated the
association between blood Pb and eGFR. Kim and Lee (2012) evaluated BLLs and eGFR (MDRD) cross-
sectionally using KNHANES (2008-2010). Compared with the lowest quartile of blood Pb (Ql:
<1.734 (ig/dL), the highest quartile (Q4: >3.010 (ig/dL) was associated with a 3.835 mL/min/1.73 m2
lower (95% CI: -5.730, -1.939) eGFR. Similarly, increased odds of a lower eGFR (<80 mL/min/1.73 m2)
were observed when comparing Q4 with Ql (OR: 1.631 (95% CI: 1.246, 2.136). In another KNHANES
(2008) study, Chung et al. (2014) evaluated blood Pb and eGFR (CKD-EPI) among adults over 20 years
old. In linear models, there was 2.61 mL/min/1.73 m2 lower (95% CI: -3.29, -1.97) eGFR for each unit
higher blood Pb. Additionally, a higher odds of reduced eGFR (<60 mL/min/1.73 m2) were observed
when comparing the highest quartile of blood Pb (Q4 mean: 4.13 (ig/dL) with the lowest (Ql mean:
1.38 (ig/dL). Buser et al. (2016) cross-sectionally evaluated the relationship between blood Pb and eGFR
(CKD-EPI) using NHANES (2007-2012). This study reported an average 2.67 mL/min/1.73 m2 lower
(95% CI: -4.78, -0.56) eGFR when comparing the highest quartile (Q4: >1.82 (ig/dL) to the lowest
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quartile (Q1 <0.79 (ig/dL) of blood Pb. Another large NHANES (2003-2014) analysis (Jain. 2019)
evaluated BLLs and decreased kidney function (eGFR (CKD-EPI) <60 mL/min/1.73 m2). Participants
with BLLs >2.15 (ig/dL had greater odds (OR: 1.567 [95% CI: 1.346, 1.823]) of lower kidney function
compared with those with lower BLLs.
As described above, Lee et al. (2020) conducted an EWAS on 262 environmental chemicals using
NHANES (1999-2016). Reduced eGFR (<60 mL/min/1.73 m2, CKD-EPI) was assessed within this study.
The discovery data set was created by combining five NHANES cycles (1999-2000, 2003-2004, 2007-
2008, 2011-2012, and 2015-2016). Individual regression analyses were conducted for each survey cycle
and combined using a random-effects meta-analysis. Identified chemicals were then reanalyzed in the rest
of the survey cycles (2001-2002, 2005-2006, 2009-2010, and 2013-2014) and referred to as the
Validation" set. Blood Pb was analyzed in both the discovery and validation sets for reduced eGFR.
Overall, the association between reduced eGFR and one standard deviation (SD) increase in the log-
transformed blood Pb concentration was positive in both the discovery (OR: 1.30 [95% CI: 1.19, 1.42])
and validation (OR: 1.20 [95% CI: 1.10, 1.30]) sets.
5.4.1.2 Toxicological Studies of Glomerular Filtration Rate
In the previous Pb ISAs, some studies found that exposure to Pb-induced changes in indicators of
renal function and structure. For example, in a few studies by the same authors in rats (mean blood Pb
-30-45 (ig/dL), decreases in GFR consistent with hyperfiltration and renal hypertrophy were reported
(U.S. EPA. 2013). This is important given that kidney hyperfiltration can be seen in early-stage diabetes,
and over time, can eventually lead to decreased kidney function. Since the publication of the document,
Shi et al. (2020) reported that a 28-day Pb drinking water exposure in rats (BLL of-10.21 (ig/dL) resulted
in a statistically significant decrease in GFR. Decreases is GFR are also important, potentially indicating
progression of kidney disease and ultimately, kidney failure. Additional information on the experimental
design of this toxicological study can be found in Table 5-5.
5.4.1.3 Integrated Summary of Glomerular Filtration Rate
The 2006 Pb AQCD (U.S. EPA, 2006) reported an association between BLLs and accelerated
decreases in eGFR in CKD patients (Yu et al., 2004). Since the 2013 Pb ISA (U.S. EPA, 2013),
longitudinal cohort studies have all reported an association between increases in BLLs and decreases in
eGFR (Chung et al., 2020; Liu et al., 2020; Harari et al., 2018; Pollack et al., 2015). In agreement with
these longitudinal studies, cross-sectional epidemiologic studies from the previous and current review
generally reported positive associations between measures of blood or bone Pb concentrations and
decreased eGFR(Jain, 2019; Buser et al„ 2016; Chung et al., 2014; Kim and Lee, 2012; Navas-Acien et
al., 2009). In agreement with the majority of the epidemiologic evidence, an animal toxicological study
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reported that Pb-exposed rats had a statistically significantly lower (p <0.05) GFR relative to control rats
(Shi et al.. 2020) (Section 5.4.1.2). When considered as a whole, there is clear evidence that exposure to
Pb can result in a decrease in eGFR.
5.4.2 Albumin, Creatinine, and Albumin-to-Creatinine Ratio
5.4.2.1 Epidemiologic Studies of Albumin, Creatinine, and Albumin-to-Creatinine
Ratio
Increased levels of creatinine in blood or serum or decreased levels of these markers in urine can
be indicative of impaired kidney function. The 2013 Pb ISA (U.S. EPA. 2013) noted positive associations
between biomarkers of Pb exposure and serum creatinine. The ISA highlighted several longitudinal NAS
studies (Tsaih et al.. 2004; Kim et al.. 1996) and cross-sectional analyses (Akesson et al.. 2005) that
evaluated the effects of bone and blood Pb exposure on creatinine. Several of these analyses indicated
positive associations between biomarkers of Pb exposure and increases in creatinine. Kim et al. (1996)
conducted a sensitivity analysis that excluded a subset of the cohort with high past Pb exposures. The
results among individuals with past Pb exposures (measured as early as 1979) <10 (ig/dL were consistent
with the results based on the entire cohort, suggesting that the association between blood Pb and increased
serum creatinine is not heavily influenced by high past Pb exposures. In addition, increases in urinary
ALB and ACR are also commonly used to assess kidney function. All of these measures can help indicate
how well the kidney is functioning. Recent evidence continues to generally indicate an increased
association between biomarkers of Pb exposure and increases in ALB, creatinine, and ACR. Study-
specific details, including BLLs, study population characteristics, confounders, and select results from
these studies are highlighted in Table 5-6. Study details in Table 5-6 include standardized results (ACR
associated with a 1 (ig/dL increase in BLL or a 10 |ig/g increase in bone Pb level) as well as results that
could not be standardized with the information provided in each paper.
The BioCycle study evaluated several different markers of kidney function, including eGFR
(Section 5.4.1.1) and blood Pb among premenopausal women (Pollack et al.. 2015). During the course of
two menstrual cycles (8 weeks) there was a 3.47% increase (95% CI: 0.85, 6.16) in creatinine with each
doubling of blood Pb. However, there was no associated increase in ALB (-0.38% [95% CI: -1.28, 0.52])
during the study period. This study also assessed several other biomarkers of kidney damage and
indicated no further associations between blood Pb and kidney dysfunction. In an NHANES (2007-2012)
analysis, Buser et al. (2016) evaluated urinary ALB. However, the study did not observe an association
with urinary ALB and blood Pb (6.29 mg/g creatinine (95% CI: -6.39, 20.80) when comparing the
highest quartile (Q4: >1.82 (ig/dL) with the lowest quartile (Q1 <0.79 (ig/dL) of blood Pb.
Muiai et al. (2019) evaluated ACR within the SPHERL study. The SPHERL study was a cross-
sectional analysis evaluating newly hired male Pb workers at battery manufacturing and Pb recycling
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plants in the United States and assessed blood Pb (taken at baseline, before large potential occupational
exposure) and concurrent ACR. The authors indicated a null association between blood Pb and ACR
(-0.071 mg/g (95% CI: -0.14, 0.59), among the men enrolled in the study. Jain (2019) also assessed
decreased kidney function as ACR >30 mg/g (measure of albuminuria) among NHANES (2003-2014)
participants. For those with BLLs >2.15 (ig/dL, increased odds (OR: 1.206 [95% CI: 1.05, 1.385]) of
ACR >30 mg/g creatinine were observed compared with those with lower BLLs. However, Zhu et al.
(2019) evaluated blood Pb and ACR within another NHANES (2009-2012) cohort and reported a null
association between quartiles of blood Pb and ACR.
In the EWAS study, described above, discovery and validation sets were created using NHANES
(1999-2016) data on 262 environmental chemicals. In addition to other indicators, the authors also
evaluated albuminuria (ACR >30 mg/g). In the discovery set, individual regression analyses were
conducted for each survey cycle and combined using a random-effects meta-analysis. Chemicals with an
FDR <1% in the meta-analysis were considered as potential risk factors for CKD and reanalyzed in the
validation set. Blood Pb was generally associated with albuminuria measured as ACR >30 mg/g, but the
results were less consistent when the discovery set was compared with the validation set (discovery set:
OR: 1.23 [95% CI: 1.07, 1.42], validation set: OR: 1.08 [95% CI: 0.97, 1.20]). However, when measured
as ACR >300 mg/g, greater congruence was observed between the estimates (discovery set: OR: 1.39
[95% CI: 1.22, 1.59], validation set: OR: 1.38 [95% CI: 1.16, 1.63]).
A small, randomized control trial (RCT) (n = 32) evaluated patients with renal insufficiency
(measured as creatinine level > 132.6 |imol/L and < 353.8 |imol/L) and mild elevated body lead burden
(>150 |ig and < 600 |ig per 72-hour urine collection). The treatment group received two months of
chelation therapy, while the control group did not. Despite similar rates of progression of renal
insufficiency at the start of trial, the chelation group had slower progression of renal insufficiency, and a
greater reduction in body lead burden, compared to the control group. In this RCT, chelation therapy with
(EDTA), resulted in a slower progression of renal insufficiency (Lin et al.. 1999). A similar RCT with 64
eligible patients evaluated chelation therapy for a longer period (weekly for 24 months), compared to a
control group. Similar to Lin et al. (1999) this trial indicated that chelation therapy resulted in improved
renal function over the course of the study (Lin et al.. 2003).
5.4.2.2 Toxicological Studies of Creatinine and Albumin
Previous Pb reviews included animal toxicological studies reporting that exposure to Pb increased
serum levels of creatinine (see Table 4-28 in the 2013 Pb ISA). For example, Berrahal et al. (2011)
reported that in rats with BLLs of 12.7 (ig/dL and 7.5 (ig/dL, serum creatinine levels were elevated. In
addition, an animal toxicology study demonstrated increased urinary ALB following exposure to Pb (BLL
of 20 (ig/dL). Studies published since the 2013 Pb ISA are presented in sections 5.4.2.2.1 and 5.4.2.2.2
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below. Moreover, additional information on the experimental design of toxicological studies of creatinine
and ALB published since the 2013 Pb ISA can be found in Table 5-7.
5.4.2.2.1 Creatinine
Since the publication of the 2013 Pb ISA, rodent toxicological studies have further demonstrated
changes in creatinine levels following Pb exposure via drinking water or gavage. Zou et al. (2015)
reported a statistically significant increase in serum levels of creatinine (BLL of 21.7 (ig/dL) following a
30-day exposure in mice relative to controls. Similarly, Laamech et al. (2016) reported a statistically
significant increase in plasma levels of creatinine in 40-day Pb-treated animals (18 (ig/dL BLL). In
addition, Shi et al. (2020) reported that a 28-day Pb exposure in rats (BLL of-10.21 (ig/dL) resulted in a
statistically significant increase in serum creatinine, as well as significantly (p <0.05) lower urine
creatinine (potentially indicating impaired kidney function). Following a single exposure in rats, (BLL of
-30 (ig/dL), Andielkovic et al. (2019) reported a small, but statistically significant increase (p <0.05) in
serum levels of creatinine. Finally, in kidney tissue, Gao et al. (2020) demonstrated a statistically
significant decrease in creatinine activity following a 4-week Pb exposure (BLL of 10.6 (ig/dL).
In addition to the drinking water and gavage studies mentioned above, a Pb nitrate nanoparticle
inhalation study reported changes in creatinine levels. Following Pb nitrate nanoparticle inhalation for
11 weeks (but not 2 or 6 weeks, BLL at 11 weeks was 8.5 (.ig/dL). Dumkova et al. (2020a) reported a
statistically significant decrease in blood creatinine in mice. Notably, this inhalation study demonstrated a
decrease in creatinine levels while the oral exposure studies mentioned above generally demonstrated an
increase in these markers. Given this is a single inhalation study, it is difficult to deduce whether the
result is repeatable, and if so, whether the difference is due to the route of exposure, the use of synthetic
Pb particles, or another factor.
Finally, not all animal toxicological studies demonstrated changes in creatinine levels. Corsetti et
al. (2017) reported no significant difference in serum creatinine levels following a 45-day Pb exposure
(BLL 21.6 (ig/dL) relative to control animals. Carlson et al. (2018) similarly reported that exposure to Pb
resulting in a BLL of 2.89 (ig/dL yielded creatinine levels that were not always within reference ranges
but were not statistically different from the levels of control mice. Furthermore, in an analysis using Pb-
oxide nanoparticles, and in contrast to their previous study (see (Dumkova et al.. 2020a) above),
(Dumkova et al.. 2020b) observed no change in creatinine levels at 2, 6, or 11 weeks, potentially
indicating a difference between Pb-oxide and Pb nitrate nanoparticle inhalation exposure (BLLs ranged
from 10.4 to 17.4 (.ig/dL). It should be noted that creatinine levels were within reference ranges in both
studies.
When the animal toxicological studies are considered together, there is evidence that exposure to
Pb can result in changes in creatinine levels. Following drinking water or gavage exposure, most studies
demonstrated an increase in serum creatinine levels, which could indicate impaired kidney function.
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However, it should be noted that a couple of these oral exposure studies, one of which was at a very low
BLL (2.89 (ig/dL), reported no change following Pb exposure. Results using Pb nanoparticle inhalation
exposure were more variable, demonstrating either a decrease or no change in creatinine levels.
5.4.2.2.2 Albumin
Since the publication of the 2013 Pb ISA, no animal toxicological studies have examined changes
in urinary ALB. Moreover, none of the existing studies demonstrated an increase in ALB serum or blood
levels. Studies either demonstrated no effect (Dumkova et al.. 2020a; Andielkovic et al.. 2019; Corsetti et
al.. 2017) or a decrease in ALB levels (Dumkova et al.. 2020b) following exposure to Pb.
5.4.2.3 Integrated Summary of Creatinine and Albumin Levels
Increased levels of creatinine in blood or serum, or a decrease in urine, can be indicative of
impaired kidney function. The 2013 ISA (U.S. EPA. 2013) included longitudinal epidemiologic studies
that evaluated the effect of bone Pb exposure on serum creatinine levels (Tsaih et al.. 2004; Kim et al..
1996). These studies both reported positive associations between increases in serum creatinine levels and
bone Pb measurements. These results are in agreement with a more recent study in premenopausal women
reporting a positive association between serum creatinine levels and increasing BLLs (Pollack et al..
2015).
Positive epidemiologic associations are supported by animal toxicological studies with BLLs
below 30 (ig/dL from both current and previous reviews. In particular, studies using oral exposures
generally demonstrate higher creatinine levels in Pb-exposed animals when compared with controls (Shi
et al.. 2020; Andielkovic et al.. 2019; Laamech et al.. 2016; Zou et al.. 2015; Berrahal et al.. 2011; Roncal
et al.. 2007) (Section 5.4.2.2). However, more recent oral exposure studies demonstrated no change in
serum creatinine levels in laboratory animals following Pb exposure (Carlson et al.. 2018; Corsetti et al..
2017).
With respect to inhalation exposure to Pb (Dumkova et al.. 2020b) reported no change in
creatinine levels. Nonetheless, it should be noted that Dumkova et al. (2020b) was unique in that it used
engineered Pb-oxide nanoparticles to expose mice via inhalation, rather than exposure through drinking
water or ingestion as in other animal toxicological studies. Moreover, these results are in contrast to those
from the same authors using Pb-nitrate nanoparticles, which reported a decrease in creatinine levels at
similar time points (Dumkova et al.. 2020a). Thus, in animal toxicological studies, it is possible that the
route of exposure or the type of Pb particles used (e.g. Pb-oxide versus Pb-nitrate) could influence serum
creatinine levels. Nonetheless, the overall evidence indicates that exposure to Pb can produce increased
levels of creatinine in blood or serum from both epidemiologic and animal toxicological studies. With
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respect to the levels of ALB, there is little evidence from epidemiologic or animal toxicological studies
that exposure to Pb can increase serum, blood, or urine ALB levels.
5.4.3 Uric Acid and Urea
5.4.3.1 Epidemiologic Studies of Uric Acid and Urea
UA is excreted in the urine and is the product of purine metabolism. Increased serum UA (SUA)
levels can be indicative of reduced kidney excretion and is associated with multiple clinical outcomes
including gout and CKD. Exposure to Pb is thought to alter UA homeostasis by effecting its kidney
excretion (Emmerson and Ravenscroft. 1975) and increased UA can result in Pb-related nephrotoxicity
(Weaver et al.. 2005). Recent epidemiologic evidence supports an association between biomarkers of Pb
exposure and increases in SUA. Study-specific details, including BLLs, study population characteristics,
confounders, and select results from these studies are highlighted in Table 5-8. Study details in Table 5-8
could not be standardized (UA associated with a 1 (ig/dL increase in blood Pb) with the information
provided in each paper.
Park and Kim (2021) evaluated the association between blood Pb and SUA levels using
KNHANES (2016-2017). This study noted higher SUA among women (0.019 mg/dL [95% CI: 0.001,
0.037 mg/dL]), but not men (-0.018 mg/dL [95% CI: -0.038, 0.002 mg/dL]) for each doubling of log-
transformed blood Pb. This study also considered hyperuricemia (SUA levels >7 mg/dL in males or
>6 mg/dL in females) but indicated null associations for both women and men. Arrebola et al. (2019)
evaluated continuous SUA levels and the presence or absence of hyperuricemia (SUA levels >7 mg/dL in
males or >6 mg/dL in females, SUA lowering medication use, or gout diagnosed by a physician) in the
BIOAMBIENT.ES study. The study population had relatively low BLLs (median: 0.106 (ig/dL). BLLs
were not associated with SUA levels (0.01 ng/dL [95% CI: -0.02, 0.04 mg/dL]) or with hyperuricemia
(OR: 1.12 [95% CI: 0.90, 1.41]). Another KNHANES analysis evaluated the effect between
hyperuricemia (SUA levels >7 mg/dL in males or >6 mg/dL in females) and BLLs (Jung et al.. 2019).
This study also did not indicate an association between blood Pb and hyperuricemia (Figure 5-5).
Notably, no epidemiologic studies have examined the potential relationship between exposure to
Pb and changes in measures of urea.
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3.5
3
2.5
2
1.5
1
0.5
0
Q1
Lead men
Q2
Q3
3=0.202
Q4
3.5
3
2.5
^ 7
1.5
1
0.5
0
Q1
Lead women
Q2
Q3
a =0,684
Q4
Adapted from: Jung et al. (2019).
Figure 5-5 Association between blood Pb and hyperuricemia among men
and women, Korea National Health and Nutrition Examination
Survey, 2016.
5.4.3.2 Animal Toxicological Studies of Uric Acid and Urea
Previous Pb reviews contained animal toxicological studies reporting that exposure to Pb
increased serum levels of creatinine (see Table 4-28 in the 2013 Pb ISA). Roncal et al. (2007)
demonstrated an increase in both serum UA and BUN following exposure to Pb (BLL 26 (.ig/dL).
Similarly, Wang et al. (2010) demonstrated increased serum urea nitrogen levels following exposure to
Pb. However, Pb levels were measured in serum, and thus the BLL was unknown. Studies published since
the 2013 Pb ISA are presented in sections 5.4.3.2.1 and 5.4.3.2.2 below. Additional information on the
experimental design of toxicological studies of UA and urea published since the 2013 Pb ISA can be
found in Table 5-10.
5.4.3.2.1 Uric Acid
Since the publication of the 2013 Pb ISA, Shi et al. (2020) reported that rats with a BLL of
-10.21 (ig/dL had a statistically significant increase in UA relative to controls. However, Laamech et al.
(2016) reported a statistically significant decrease (18 (ig/dL BLL), and Andielkovic et al. (2019) reported
no change (-30 (ig/dL BLL) in UA following exposure to Pb. Thus, there is limited evidence from animal
toxicologic studies for increased levels of UA.
5.4.3.2.2 Urea
Since the publication of the 2013 Pb ISA, rodent toxicological studies have further demonstrated
changes in measures of urea following Pb exposure via drinking water or gavage. Zou et al. (2015)
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reported a statistically significant increase in BUN (BLL of 21.7 (ig/dL) following a 30-day exposure in
mice relative to controls. Similarly, following exposure to Pb.Laamech et al. (2016) reported a
statistically significant increase in plasma levels of urea (18 (ig/dL BLL). In addition, Shi et al. (2020)
reported that a 28-day Pb exposure in rats (BLL of-10.21 (ig/dL) resulted in a statistically significant
increase in BUN. Similarly, in kidney tissue, Gao et al. (2020) demonstrated a statistically significant
increase in BUN activity following a 4-week Pb exposure (BLL of 10.6 (.ig/dL). In contrast to studies that
found an increase in serum BUN following Pb exposure, Andielkovic et al. (2019) reported a statistically
significant (p <0.05) decrease in serum BUN (BLL of -30 (ig/dL). In addition, both Corsetti et al. (2017)
(BLL 21.6 (ig/dL) and Carlson et al. (2018) (BLL of 2.89 (ig/dL) reported that exposure to Pb did not
result in urea levels that were statistically different from those of control animals.
In addition to the studies above, a couple of Pb nanoparticle inhalation studies (by the same
authors) reported mixed results. Following Pb nitrate nanoparticle inhalation for 11 weeks (but not 2 or
6 weeks; BLL at 11 weeks was 8.5 (.ig/dL). Dumkova et al. (2020a) reported a statistically significant
decrease in urea levels. However, in an analysis using Pb-oxide nanoparticles, no change in urea was
reported at 2, 6, or 11 weeks (Dumkova et al.. 2020b). potentially indicating a difference between Pb-
oxide and Pb nitrate nanoparticle inhalation exposure (BLLs ranged from 10.4 to 17.4 (.ig/dL).
The majority of the studies published since the last ISA indicate that oral exposure to Pb can
result in changes in measures of urea (e.g. BUN). Most of these studies demonstrated an increase in serum
urea levels, consistent with impaired kidney function. Inhalation studies conducted by the same laboratory
were more variable, demonstrating no change or decreases in these markers. Additional information
regarding the experimental designs of the of urea and UA studies included in this section can be found in
Table 5-9.
5.4.3.3 Integrated Summary of Uric Acid and Urea
Similar to other molecular markers that estimate kidney function, increased SUA, urea, and BUN
levels can be indicative of impaired kidney function. Increased SUA levels are also associated with
multiple clinical outcomes including gout and CKD. In an epidemiologic study, Park and Kim (2021)
reported an increase in SUA among women, but not men. However, Arrebola et al. (2019) and Jung et al.
(2019) reported that BLLs were not associated with either SUA levels or the presence of hyperuricemia.
Animal toxicology studies were similarly mixed. Thus, there is limited evidence from epidemiologic and
animal toxicological studies for increases in the levels of UA following Pb exposure.
With respect to measures of urea, some animal toxicological studies with mean blood Pb values
<30 (ig/dL have demonstrated a relationship between exposure to lead and increased serum BUN levels
(Shi et al.. 2020; Laamech et al.. 2016; Zou et al.. 2015). Similarly, Gao et al. (2020) demonstrated a
statistically significant increase in kidney tissue BUN levels. Other oral exposure studies were mixed,
either showing a decrease (Andielkovic et al.. 2019) or no effect (Carlson et al.. 2018; Corsetti et al..
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2017). Inhalation studies by the same authors were also mixed, with some studies demonstrating a
significant decrease in blood urea levels following inhalation of engineered Pb-nitrate (Dumkova et al..
2020b) but not Pb-oxide n an o p art i c 1 c s (D u m k o v a et al.. 2020a) in rats. Some evidence from animal
toxicology studies suggests that oral exposure can increase the levels of urea in blood following exposure
to Pb.
5.4.4 Proteinuria and Hematuria
5.4.4.1 Epidemiologic Studies of Proteinuria and Hematuria
Increased levels of protein (proteinuria) and blood cells (hematuria) in the urine can be markers
of renal damage. Hematuria can either be benign or indicative of more serious outcomes including
glomerulonephritis, CKD, kidney stones, or cancer. The 2013 Pb ISA (U.S. EPA. 2013) did not include
any epidemiologic studies of proteinuria or hematuria. Study-specific details, including BLLs, study
population characteristics, confounders, and select results from more recent studies examining these
endpoints are highlighted in Table 5-10. Study details in Table 5-10 include standardized results
(associated with a 1 (ig/dL increase in BLL) as well as results that could not be standardized with the
information provided in each paper.
Chung et al. (2014) evaluated the association between blood Pb and proteinuria using KNHANES
(2008). Proteinuria was defined as >1 on a urine dipstick test (equivalent to >30 mg/dL). This study
indicated that when compared with the lowest quartile (mean: 1.38 (.ig/dL). the odds of proteinuria (OR:
1.22 [95% CI: 1.00, 1.50]) were higher among participants in the highest quartile (mean: 4.13 (ig/dL).
Han et al. (2013) evaluated the association between hematuria (>1 on urine dipstick test) and BLLs using
KNHANES (2008-2010). A null association was observed when the highest quartile (Q4 >3.22 (ig/dL)
was compared with the lowest (Ql: <1.89 (ig/dL) (OR: 0.78 [95% CI: 0.443, 1.361]).
5.4.4.2 Toxicological Studies of Proteinuria and Hematuria
The previous ISA contained no evidence of proteinuria or hematuria from animal toxicological
studies with reported BLLs. One study reported an increase in urinary protein levels but only measured
Pb in serum Wang et al. (2010). No animal toxicological studies have been conducted with BLLs
examining these outcomes since the 2013 Pb ISA. Thus, consistent with the epidemiologic studies
presented above, there is only limited evidence for an effect of Pb on proteinuria and no evidence for an
effect on hematuria.
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5.4.4.3 Integrated Summary of Proteinuria and Hematuria
There is little evidence from epidemiologic or animal toxicological studies that exposure to Pb
results in proteinuria or hematuria
5.4.5 N-Acetyl-p-D-Glucosaminidase and ^-Microglobulin
5.4.5.1 Epidemiologic Studies of N-Acetyl-P-D-Glucosaminidase and P2-Microglobulin
Many markers of kidney dysfunction may be insensitive for early detection of kidney damage.
Recently, the development of early biological effect (EBE) markers of preclinical kidney damage has
received substantial attention. Exposure to Pb is thought to directly affect the deterioration of tubular
function, which can lead to the loss of essential divalent metals. The renal tubular biomarker N-acetyl-|3-
D-glucosaminidase (NAG) is a lysosomal enzyme that is sensitive to renal impairment. Another renal
tubular biomarker, |32-microglobulin (P2-MG), is typically reabsorbed through glomerular filtration.
Increases in either NAG or P2-MG correspond to damage to the renal tubules. Study-specific details,
including BLLs, study population characteristics, confounders, and select results from these studies are
highlighted in Table 5-11. Study details in Table 5-11 could not be standardized (associated with a
1 (ig/dL increase in blood Pb) with the information provided in each paper.
Lim et al. (2016) evaluated the association between BLLs and both NAG and P2-MG in the
Korean Research Project on the Integrated Exposure Assessment to Hazardous Materials for Food Safety
(KRIEFS). This study indicated null associations between log-transformed blood Pb and both NAG (0.09
units/g creatinine [95% CI: -0.05, 0.23 units/g creatinine]) and P2-MG (0.01 |ig/g creatinine [95% CI:
-0.13,0.15 |ig/g creatinine]). Jung et al. (2016) also evaluated the association between blood Pb and
NAG among participants residing near a cement plant in South Korea. There were null associations when
high NAG levels (>5.67 U/L) were compared with low NAG levels between quartiles of blood Pb and
NAG.
5.4.5.2 Toxicological Studies of N-Acetyl-P-D-Glucosaminidase and P2-Microglobulin
A study from the 2013 Pb ISA indicated an increase in P-2 microglobulin and N-acetyl-P-D-
glucosaminidase following Pb exposure (Wang et al.. 2010). However, this study only measured Pb levels
in serum (serum Pb level: 20 (ig/dL) and thus, the BLL is unknown. Similarly, Javakumar et al. (2009)and
Khalil-Manesh et al. (1992b) reported a change in N-acetyl-P-D-glucosaminidase following Pb exposure
(BLLs and 45 (ig/dL, and >55 (ig/dL, respectively). Since the 2013 Pb ISA, no animal toxicological
studies have been conducted with BLLs less than 30 (ig/dL to examine these markers.
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5.4.5.3 Integrated Summary of N-Acetyl-P-D-Glucosaminidase and P2-Microglobulin
Few epidemiologic studies have been conducted examining |3-2 microglobulin and N-acetyl-|3-D-
glucosaminidase following Pb exposure, and these studies reported no association with BLLs. With
respect to animal toxicological studies, a few studies from previous reviews demonstrated changes in N-
acetyl-|3-D-glucosaminidase following Pb exposure, but a couple of these studies were at BLLs
>55 (ig/dL. Thus, when considered together, epidemiologic and animal toxicological studies provide little
evidence for an effect of Pb exposure on these markers at BLLs <30 (ig/dL.
5.4.6 Toxicological Studies of Other Indicators of Kidney Function
In addition to the markers potentially indicating impaired kidney function discussed above, other
markers have been examined in a small number of studies. Increases in total serum protein can also be
indicative of impaired kidney function. However, Andielkovic et al. (2019) reported no change in total
serum protein following Pb exposure (BLL -30 (ig/dL). Moreover, other studies either reported no
changes or decreases in total protein in blood at timepoints ranging from 2 weeks to 11 weeks following
Pb nitrate (Dumkova et al.. 2020a) or Pb-oxide (Dumkova et al.. 2020b) nanoparticle inhalation exposure
(BLLs <17.4 (ig/dL in these studies).
Changes in the balance of metal ions in the kidney and blood can also be indicative of impaired
kidney function. In particular, lower calcium levels can be indicative of kidney disease. Dumkova et al.
(2020a) reported a significant decrease in calcium levels in the kidney but not in blood following Pb
nitrate nanoparticle inhalation for 2 weeks (but not 6 or 11 weeks when BLLs were higher; BLL at
2 weeks was 4 (ig/dL). No changes in the blood levels of sodium or potassium were reported, but there
was a statistically significant decrease in phosphorous levels in blood at 2 and 11 weeks (but not at
6 weeks). In an additional analysis using Pb-oxide nanoparticles, Dumkova et al. (2020b) reported a
statistically significant decrease in kidney calcium levels after 2 and 6 weeks, but not after 11 weeks of
exposure (BLLs: 10.4 (ig/dL at 2 weeks, 14.8 (ig/dL at 6 weeks and 17.4 (ig/dL at 11 weeks). In addition,
this study found no changes in sodium or potassium levels in the kidney at any time point. There were
also no changes in calcium, potassium, or sodium levels in the blood. Moreover, Andielkovic et al. (2019)
reported: 1) a statistically significant decrease in serum calcium and iron; 2) no change in blood copper,
zinc, or phosphorus levels; and 3) a decrease (p <0.05) in kidney tissue zinc, but not copper following Pb
exposure (BLL -30 (ig/dL). Finally, Zou et al. (2015) reported no change in zinc levels but a decrease in
iron levels in blood relative to control animals. When considered as a whole, there is limited evidence for
changes in calcium and other ion levels in blood or tissue following exposure to Pb. Additional
information on the experimental design of toxicological studies presented in this section can be found in
Table 5-12.
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5.5
Toxicological Studies of Metal Co-Exposures with Pb
A limited number of studies evaluated the effect of Pb on the kidney in conjunction with exposure
to other metals. Andielkovic et al. (2019) evaluated the effect of Pb exposure in combination with
cadmium. Although the levels of creatinine, BUN, and UA were similar following co-exposure with
cadmium, total serum protein and ALB levels were statistically lower than controls following co-
exposure. In an additional study, Zou et al. (2015) reported a statistically significant increase in serum
levels of creatinine and BUN following co-exposure of Pb and zinc, but the levels of these markers were
lower than the levels following exposure to Pb alone.
With respect to metal ions and co-exposure, Andielkovic et al. (2019) reported that co-exposure
of Pb with cadmium resulted in a statistically significant decrease (p <0.05) in the levels of zinc (but did
not exacerbate the decrease compared with Pb alone) and no change in copper ion levels (similar to Pb
alone) in kidney tissue. However, co-exposure with cadmium did result in a greater decrease in serum
calcium, iron, and blood copper levels, but not zinc blood levels when compared with exposure to Pb
alone. In addition, Zou et al. (2015) reported that co-exposure with zinc significantly increased blood iron
levels relative to Pb exposure alone.
Overall, only a few studies have examined the potential effects of metal co-exposure on kidney-
related endpoints. Moreover, these studies varied in their co-exposure metals and outcome assessments.
Thus, it is difficult to draw conclusions on the effects of metal co-exposure with Pb on either markers of
kidney function or ion concentrations.
5.6 Activation of Renin-Angiotensin-Aldosterone System
The renin-angiotensin-aldosterone system (RAAS) plays an important role in the regulation of
blood pressure and kidney homeostasis. For example, angiotensin II (Ang II) stimulates arteriolar
vasoconstriction, leading to increases in blood pressure or hypertension. Angiotensin-converting enzyme
(ACE) is involved in the activation of Ang II. The 2013 Pb ISA stated that vascular reactivity to Ang II
increased following Pb exposure (Robles et al.. 2007). In addition, exposure to Pb resulted in increases in
kidney or serum ACE activity and renal Ang II-positive cells (Rodriguez-Iturbe et al.. 2005; Sharifi et al..
2004; Carmignani et al.. 1999). Moreover, use of an ACE inhibitor or blocking the Ang II receptor type 1
(AT-1) ameliorated Pb-induced increases in blood pressure (SimSes et al.. 2011). Since the 2013 Pb ISA,
Fioresi et al. (2014) reported no change in ACE activity in plasma and cardiac tissue. Taken together,
there is some evidence from older studies to suggest that exposure to Pb can result in changes in RAAS.
Additional information on the study design of Fioresi et al. (2014) can be found in Table 5-12.
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5.7
Renal Outcomes Among Children
The 2013 Pb ISA (U.S. EPA. 2013) and 2006 Pb AQCD (U.S. EPA. 2006) highlighted several
studies indicating a lack of association between biomarkers of Pb exposure and renal outcomes among
children. Many studies presented previously were among children with high exposures to Pb. Fadrowski
et al. (2010) conducted an NHANES analysis that evaluated relatively low blood Pb values (median:
1.5 (ig/dL) and two different measures of eGFR (cystatin C-based and creatinine-based). This study
indicated higher eGFR based on an association between cystatin C and the highest quartile (>2.6 (ig/dL)
compared with the lowest (<1 (ig/dL). More recent analyses not only continue to evaluate children with
low BLLs, but also use techniques to more accurately measure GFR (either directly or an estimate) in
children. Study-specific details, including Pb biomarker levels, study population characteristics,
confounders, and select results from these studies are highlighted in Table 5-13. Study details in
Table 5-13 include standardized results (associated with a 1 (ig/dL increase in BLL) as well as results that
could not be standardized with the information provided in each paper.
A recent longitudinal analysis evaluated the association between the erythrocyte fraction of Pb
(Ery-Pb) in maternal blood and subsequent measurements of renal function, including kidney volume,
eGFR (calculated based on serum cystatin C and deemed appropriate for use in children), and serum
cystatin C, among children (~ 4.5 years) (Skroder et al„ 2016). The Ery-Pb was assessed at both 14 weeks
(GW14) and 30 weeks (GW30) of gestation. Linear regression analyses identified an association between
decreased kidney volume and maternal Ery-Pb at 30 weeks of gestation (-0.071 cm3/m2 [95% CI: -1.4,
-0.030]), but not at 14 weeks of gestation (-0.061 cm3/m2 [95% CI: -0.36, 0.24]). When stratified by sex,
this association was stronger among girls (-1.1 cm3/m2 [95% CI: -2.1, -0.049]) than among boys
(-0.80 cm3/m2 [95% CI: -1.80, 0.20]), for each 10 (ig/kg increase in Ery-Pb. However, no differences in
effect were observed when this outcome was stratified by birthweight or by children with stunted height.
When considering other markers of renal dysfunction, no associations were present for eGFR (GW14
0.089 mL/min/1.73 m2 [95% CI: -0.012, 0.30]; GW30 0.71 mL/min/1.73 m2 [95% CI: -0.24,0.17]) or
serum cystatin C (GW14 -0.00088 mg/L [95% CI: -0.0028, 0.001]; GW30 0.000027 [95% CI: -0.0018,
0.0018]).
Fadrowski et al. (2013) conducted a cross-sectional study evaluating children (aged 1-16) with
CKD who were part of the Chronic Kidney Disease in Children (CKiD) prospective study. This study
measured GFR directly by measuring the plasma disappearance inhexol curves (children had blood draws
at 10, 20, 120, and 300 minutes after an injection of inhexol). The average percent change in GFR within
the study was -2.1% (95% CI: -6.0, 1.8) for a 1 (ig/dL increase in blood Pb. In the pediatric population,
there are two main diagnoses for CKD: glomerular and nonglomerular. Glomerular CKD diagnoses
include focal segmental glomerulosclerosis, hemolytic uremic syndrome, and systemic immunological
diseases (systemic lupus erythematosus), whereas nonglomerular CKD includes
aplastic/hypoplastic/dysplastic kidneys, reflux nephropathy, obstructive uropathy, and congenital urologic
disease. Generally, nonglomerular CKD has an earlier onset and a slower rate of disease progression
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(Hooper et al.. 2021). When stratified by the type of CKD (glomerular versus nonglomerular), children
with glomerular CKD experienced a -12.1% change (95% CI: -22.2, -1.9) in GFR, compared with a
-0.7% change (95% CI: -4.8, 3.4) among those with nonglomerular CKD. In another cross-sectional
analysis, Cardenas-Gonzalez et al. (2016) evaluated BLLs and two biomarkers of kidney injury (Kidney
Injury Molecule 1 [KIM-1] and neutrophil gelatinase-associated lipocalin [NGAL]) among Mexican
children living in an area with a high prevalence of CKD. This study indicated null associations between
blood Pb and biomarkers of kidney injury (results not shown).
An NHANES (1999-2006) analysis evaluated blood Pb and SUA among adolescents aged 12-19
(Hu et al.. 2019). This study considered several confounders related to sociodemographic factors, blood
biochemistry markers, and dietary intake. Overall, a one-unit increase in natural log (ln)-transformed
blood Pb was associated with a 0.14 mg/dL (95% CI: 0.10, 0.17) higher SUA. Additionally, the
magnitude of the association was larger when examining elevated SUA (>5.5 mg/dL) and a one-unit
increase in ln-transformed blood Pb (OR: 1.29 [95% CI: 1.17, 1.42]). Moreover, a restricted cubic spline
analysis indicated a linear dose-response relationship between ln-transformed blood Pb and both
continuous SUA and elevated SUA (>5.5 mg/dL) (Figure 5-6). Additionally, Hu et al. (2019) evaluated
several of the model covariates (e.g. sex, race, and eGFR) as effect measure modifiers. The comparisons
for these are shown in Figure 5-7. The authors reported that there were generally no modifications
between blood Pb and other adjusted variables, except for educational attainment. Thus, the positive
association remained, regardless of subgrouping.
LtiBLL (ng/dL)
LnBLL (ji&'dL)
BLL = blood lead level; dL = deciliter; In = natural log; OR = odds ratio.
Source: Hu et al. (2019).
Figure 5-6 Associations between natural log blood Pb (0-4 [jg/dL) and serum
uric acid and elevated serum uric acid (>5.5 mg/g).
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Subgroups
N Mean + SD
P ( 95% CI)
P for interaction
Sex
0.056
Male
4184
5.6 ±1.2
0.11 (0.06, 0.17)
Female
4119
*-
1+
©
>-•-1
0.14 (0.09, 0.19)
Age, years
0.056
< 17
6261
4.9 ± 1.2
•-H
0.15(0.11, 0.19)
> 17
2042
5.2 ± 1.3 *¦
—¦—1
0.05 (-0.02, 0.13)
Race
0.808
Non-Hispanic White
2109
5.1 ± 1.3
0.12 (0.04, 0.20)
Nuii-fiibpajiic Black
2641
4.9 ±1.2
0.16 (0.10, 0.23)
Mexican American
2905
5.0 ± 1.3
i ¦ i
0.14 (0.08,0.19)
Other Hispanic
318
5.0 ± 1.2 ¦
0.07 (-0.12, 0.26)
Other race
330
r -) I 1 1 , ¦
-0.05 (-0.28, 0.17)
Education
0.002
< high school
6998
5.0 ± 1.3
l-»H
0.16 (0.12,0.20)
s high school
1301
5.2 ± 1.3 i 1
-0.03 (-0.12, 0.06)
Physical Activity
0.268
Sedentary
572
4.9 ± 1.4 ¦—
0.04 (-0.08, 0.16)
Low
803
5.0 ± 1.3
i m 1
0.16 (0.03, 0.28)
Moderate
584
5.0 ± 1.3 >-
—¦ 1
0.10 (-0.04, 0.24)
High
1537
5.3 ± 1.3
i—¦—i
0.14 (0.05, 0.22)
BMI, kg/m2
0.271
Tertile 1 (< 17.5)
603
4.3 ± 1.1 <—
0.06 (-0.08, 0.19)
Tertile 2 (17.5-22.1)
3291
4.7 ± 1.1
i ¦ i
0.15 (0.10, 0.20)
Tertile 3 022.1)
4333
5.3 ±1.3
-¦-i
0-07(0.01,0.12)
Serum Colinine, ng/mL
0.310
< 0.1
4014
4.8 ±1.2
0.13 (0.08,0.18)
0.1-10
3117
5.0 ± 1.3
—1
0.08 (0.02, 0.14)
£ 10
1108
5.4 +1.3 i-
0.10 (-0.02, 0.21)
eGFR, inL/iniii per 1.73 m2
0.047
Tertile 1 (< 130)
2750
4.9 ± 1.2
¦ ¦ ¦
0.19 (0.11,0.27)
Tertile 2 (130-152)
2749
5.1 + 1.3
n-
0.11 (0.05, 0.16)
Tertile 3 (> 152)
2757
5.0 ± 1.3
0.06(0.00,0.12)
"i—i—i—i—i—i—i—i—r
¦0.4 -0,3 -0,2 -0,1 0 0,1 0.2 0,3 0,4
BMI = body mass index; CI = confidence interval; eGFR = estimated glomerular filtration rate; kg = kilograms; m = meters;
min = minute; mL = milliliter; ng = nanograms; SD = standard deviation.
Source: Hu et al. (2019).
Figure 5-7 Effect measure modification between blood Pb and serum uric
acid among adolescents, National Health and Nutrition
Examination Survey 1999-2006.
5.7.1 Summary of Renal Outcomes Among Children
Skroder et al. (2016) conducted a longitudinal analysis evaluating the association between the
erythrocyte fraction of Pb (Ery-Pb) in maternal blood and subsequent measurements of renal function
among children (- 4.5 years). The Ery-Pb was assessed at both 14 weeks (GW14) and 30 weeks (GW30)
of gestation. Linear regression analyses identified an association between decreased kidney volume and
maternal Ery-Pb at 30 weeks of gestation, but not at 14 weeks. No associations were present with eGFR
or serum cystatin C. In addition, an NHANES (1999-2006) analysis evaluated blood Pb and serum SUA
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among adolescents aged 12-19 taking into account several confounders related to sociodemographic
factors, blood biochemistry markers, and dietary intake. Overall, there was a positive association between
a one-unit increase in transformed blood Pb and continuous and elevated SUA (Hu et al.. 2019). This
study also evaluated several of the model covariates (e.g. sex, race, and eGFR) in a subgroup analysis,
and no interaction was reported between blood Pb and other adjusted variables, except for educational
attainment. In addition to these studies, a cross-sectional study evaluated children (aged 1-16) with CKD
and measured GFR directly by measuring the plasma disappearance inhexol curves. Overall, this study
did not indicate an association between blood Pb and GFR, except among those with a specific type
(glomerular) of CKD (Fadrowski et al.. 2013). Similarly, an additional cross-sectional analysis did not
report an association between BLLs and biomarkers of kidney function among Mexican children living in
an area with a high prevalence of CKD (Cardenas-Gonzalez et al.. 2016). Taken together, there is limited
evidence for an effect between biomarkers of Pb exposure and renal outcomes among children.
5.8 Reverse Causality
In observational research, reverse causality occurs when an association between an exposure and
outcome is explained by the outcome that causes or alters the exposure. Reverse causality is a potential
concern in studies of kidney function due to the role of the renal system in the excretion of toxins from
the blood. Specifically, increased BLLs could result from reduced excretion due to kidney damage rather
than as a causative factor for kidney impairment. The potential for reverse causality in epidemiologic
studies is especially plausible in cross-sectional studies and studies conducted in study populations that
are already experiencing renal dysfunction. In contrast, prospective analyses that include baseline
measurements of biomarkers of Pb exposure and incident changes in renal function may help control for
the possibility of reverse causality.
The 2006 Pb AQCD (U.S. EPA, 2006) presented a longitudinal NAS study by Kim et al. (1996)
where positive associations between BLLs and serum creatinine were reported over most of the range of
serum creatinine (Figure 5-8). In locally weighted regression models, these associations were observed
within the normal creatinine range, where reduced excretion of Pb is a less likely explanation of the
observed association. A follow-up to this study evaluated the association between blood and bone Pb
levels and serum creatinine among those with serum creatinine <1.5 mg/dL (Tsaih et al., 2004). This
study indicated that the longitudinal associations did not materially change, suggesting that Pb dose
contributed to renal dysfunction.
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-~f 130
5 C 47)
o
E
A
®"
c
c
«
0)
6
E
3
k_
a)
CO
3
«
3
"O
<
115
(1.30)
100
(1.13)
85
(0.96)
"T
i 1
1 2 3
(20.7) (41.4) (62.2)
Adjusted Blood Lead, nmol/L (ng/dL)
Source: Kim et al. (1996).
130
(147)
115
(1 30)
100
(1.13)
85
(0.96)
1 1 1 1 1—
0 0.1 0.3 0.9 2.7
(2.1) (6.2) (18.6) (55.9)
Adjusted Blood Lead, jxmol/L (ng/dL)
Figure 5-8 Locally weighted smoothing plot of adjusted associations
between blood Pb levels (with [left panel] and without [right
paned] logarithmic transformation) and serum creatinine.
The use of eGFR provides a better estimate of progressive changes in renal function than
creatinine clearance. In a longitudinal study evaluated in the 2013 Pb ISA, Yu et al. (2004) indicated that
baseline BLLs were associated with a decline in eGFR among CKD patients. More recent longitudinal
analyses assessed changes in eGFR (Chung et al.. 2020; Liu et al.. 2020; Harari et al.. 2018; Pollack et al..
2015) among a variety of populations free of kidney disease at baseline. Notably, in a population-based
cohort study with an extensive follow-up period (Baseline: 1991-1994, Follow-up: 2007-2012), Harari et
al. (2018) reported that increased baseline BLLs were associated with substantial decreases in eGFR from
baseline. Since this study also adjusted for baseline eGFR, the larger decreases in kidney function
observed in participants with higher Pb exposures ostensibly occurred in participants with similar baseline
kidney function. Smaller cohort studies further supported this study by noting decreases in eGFR with
increased BLLs (Chung et al.. 2020; Liu et al.. 2020; Pollack et al.. 2015). Studies of these smaller
cohorts, with relatively short-term follow-up (Pollack et al.. 2015). cannot by themselves rule out reverse-
causality. However, when combined with larger and more robust studies of those without underlying
kidney disease at baseline (Chung et al.. 2020; Liu et al.. 2020; Harari et al.. 2018; Pollack et al.. 2015).
the smaller studies can contribute to reducing this uncertainty in the broader body of evidence.
Furthermore, several recent epidemiologic studies evaluated the association between BLLs and
the development of CKD or ESRD. In a population-based cohort in Sweden that showed Pb-related
reductions in eGFR, Harari et al. (2018) also observed a relationship between BLLs at baseline and
incident CKD after further adjustment for baseline eGFR. Additionally, a comprehensive analysis by
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Sommar et al. (2013) involved a combination of several existing cohort studies and subsequently linked
incident ESRD cases to members of the cohorts. This study identified a modest association between BLLs
and incident ESRD. These studies provide further evidence that links baseline blood Pb data to the
development of long-term kidney disease.
In addition to the epidemiologic evidence, the expanded literature base of animal toxicological
studies provides strong support that the associations reported in epidemiologic studies are the result of
exposure to Pb, not reverse causality. This is due to the large amount of evidence from animal
toxicological studies demonstrating health effects such as impaired kidney function and kidney damage
providing additional support that associations reported in epidemiologic studies are indeed the result of
exposure to Pb.
Overall, recent evidence further supports that reverse causality does not contribute substantially
to the association between higher BLLs and decreases in kidney function. Several recent studies
longitudinally evaluated either the change in eGFR from baseline or the development of CKD or ESRD
and baseline blood Pb measurements taken years prior to the assessment of kidney function. While
reverse causality may contribute to some associations between biomarkers of Pb exposure and renal
function, recent evidence does not support reverse causality as the driving force behind these associations.
5.8.1 Summary of Reverse Causality
Epidemiologic evidence has generally reported increased associations between biomarkers of Pb
exposure and renal effects, without evidence of reverse causality. Specifically, longitudinal studies
evaluating a decline in eGFR in relation to blood Pb further suggest that reverse causality does not
substantially affect the association between biomarkers of Pb exposure and decreased kidney function.
The 2006 Pb AQCD (U.S. EPA, 2006) reported an association between baseline BLLs and accelerated
decreases in eGFR in CKD patients (Yu et al., 2004). Several recent longitudinal studies among healthy
populations, free of kidney disease, also further support changes in eGFR from baseline, associated with
baseline blood Pb (Chung et al.. 2020; Liu et al.. 2020; Harari et al.. 2018; Pollack et al.. 2015)).
Specifically, Harari et al. (2018). which had an extensive follow-up period (-16 years of follow-up),
noted that increased baseline BLLs were associated with substantial decreases in eGFR from baseline.
In addition, several recent epidemiologic studies also evaluated the association between
biomarkers of Pb exposure and the development of CKD or ESRD. In the population-based cohort in
Sweden that also noted Pb-related reductions in eGFR, Harari et al. (2018) observed a relationship
between incident CKD and BLLs at baseline, after further adjustment for baseline eGFR. Additionally,
Sommar et al. (2013) combined several existing cohort studies and subsequently linked them to an ESRD
database. This study identified a modest association between BLLs and incident ESRD. These studies
provide further evidence that links baseline blood Pb data to the development of long-term kidney
disease.
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Toxicological evidence indicating associations between blood Pb and markers of oxidative stress
and impaired kidney damage provides additional support that associations reported in epidemiologic
studies are in fact the result of exposure to Pb. The combined toxicological and epidemiologic evidence
suggests that reverse causality does not substantially contribute to the association between higher BLLs
and decreased kidney function. While reverse causality may contribute to some associations between
biomarkers of Pb exposure and renal function, the available evidence does not support it as the driving
force behind these associations.
5.9 Biological Plausibility
Sections 5.3 to 5.8 of this appendix describe the health effects associated with exposure to Pb
from epidemiologic and animal toxicological studies. Based largely on the animal toxicological evidence
presented in these sections, as well as in previous ISAs and AQCDs, this section describes the biological
pathways that potentially underlie the renal outcomes identified in epidemiologic studies and that are
associated with Pb exposure. Figure 5-9 graphically depicts these proposed pathways as a continuum of
pathophysiological responses—connected by arrows—that may ultimately lead to the apical renal events
associated with exposures to Pb at concentrations observed in epidemiologic studies. Note that the role of
biological plausibility in contributing to the weight-of-evidence causality determinations reached in the
current Pb ISA is discussed in Section 5.10.
When considering the available health evidence, plausible pathways connecting Pb exposure to
the apical events reported in epidemiologic studies are presented in Figure 5-9. The first pathway begins
with oxidative stress directly resulting in kidney damage and increases in blood pressure. The second
pathway involves Pb activation of RAAS resulting in increases in blood pressure. Once these pathways
are initiated, there is evidence from experimental and observational studies that exposure to Pb may result
in a series of pathophysiological responses that could lead to adverse renal events such as CKD and
kidney failure.
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>
Pb Exposure
Chronic Kidney
+. Disease and its
Progression
Note: The boxes above represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and
the arrows indicate a proposed relationship between those effects. Solid arrows denote evidence of essentiality as provided, for
example, by an inhibitor of the pathway or a genetic knockout model used in an experimental study involving Pb exposure Shading
around multiple boxes is used to denote a grouping of these effects. Arrows may connect individual boxes, groupings of boxes, and
individual boxes within groupings of boxes. Progression of effects is generally depicted from left to right and color-coded (gray,
exposure; green, initial effect; blue, intermediate effect; orange, effect at the population level or a key clinical effect). Here,
population level effects generally reflect the results of epidemiologic studies. When there are gaps in the evidence, there are
complementary gaps in the figure and the accompanying text below.
It has been well established that exposure to Pb can stimulate the production of reactive oxygen
species and markers of inflammation in the blood or kidneys (see (U.S. EPA. 2013)). and evidence
published since the last Pb ISA further supports these findings. For example, in rats Andielkovic et al.
(2019) reported a statistically significant increase (p <0.05) in total oxidative status and the oxidative
stress index in blood following Pb exposure (-30 (ig/dL BLL). These authors also reported a decrease
(p <0.05) in the total antioxidative status in blood following Pb exposure (-30 (ig/dL BLL). Moreover, Pb
exposure to rat primary proximal tubular cells increased intracellular reactive oxygen species production
in a concentration-dependent manner (Wang et al.. 2011). In both of these studies, the authors reported
higher levels of lipid peroxidation (e.g. malondialdehyde or thiobarbituric acid reactive substance levels)
in kidney tissue (Andielkovic et al.. 2019) and primary cells (Wang et al.. 2011) relative to controls. Other
studies similarly demonstrated increased indicators of lipid peroxidation in serum and renal tissue after
exposure to Pb (Gao et al.. 2020; Shi et al.. 2020; Li et al.. 2017; Laamech et al.. 2016; Berrahal et al..
2011; Lodi et al.. 2011; Moneim et al.. 2011; Wang et al.. 2011; Masso-Gonzalez and Antonio-Garcia.
2009). This is important given that lipid peroxidation can be an indicator of tissue damage and because
numerous studies that included kidney histology have demonstrated abnormalities and damage to kidney
cells or tissue following Pb exposure (Gao et al.. 2020; Shi et al.. 2020; Alcaraz-Contreras et al.. 2016;
Laamech et al.. 2016; Basgen and Sobin. 2014; Roncal et al.. 2007; Rodriguez-Iturbe et al.. 2005; Fowler
Figure 5-9 Potential biological pathways for renal effects following Pb
exposure.
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et al.. 1980). Some of these Pb-induced kidney changes have been found to be the result of Pb-induced
cellular necrosis (Fowler et al.. 1980) or apoptosis (Rana. 2008). and studies have demonstrated that
inhibiting Pb-induced oxidative stress and inflammation can ameliorate kidney damage (Rana et al.. 2020;
Shafiekhani et al.. 2019). These kidney abnormalities could plausibly result in impaired kidney function.
Following exposure to Pb, markers of impaired kidney function such as increased levels of creatinine and
BUN) have been reported in animal toxicological studies (Shi et al.. 2020; Andielkovic et al.. 2019;
Laamech et al.. 2016; Zou et al.. 2015; Berrahal et al.. 2011; Roncal et al.. 2007). In addition, the previous
ISA included studies in which exposure to Pb resulted in either decreased (Shi et al.. 2020) or elevated
glomerular filtration rates (GFR) (Khalil-Manesh et al.. 1993; Khalil-Manesh et al.. 1992b; Khalil-
Manesh et al.. 1992a). both of which can be indicative of kidney disease. These studies demonstrated that
decreased GFR can be indicative of reduced blood filtration by the kidneys, while increased GFR can be
consistent with the hyperfiltration and renal hypertrophy that can occur in advanced diabetes.
Pb-induced oxidative stress can also lead to the adverse kidney outcomes reported in
epidemiologic studies through hypertension. As detailed in the cardiovascular disease appendix, oxidative
stress can lead to increases in blood pressure through a number of different pathways. An increase in
blood pressure due to Pb-induced oxidative stress is supported by a study demonstrating that in rats, the
antioxidant vitamin E could attenuate both Pb-induced oxidative stress and blood pressure increases
(Vaziri et al.. 1999). This is important given that a chronic increase in blood pressure can lead to
glomerular and renal vasculature injury, which could plausibly result in renal dysfunction and CKD.
The second pathway by which exposure to Pb could potentially lead to the outcomes reported in
epidemiologic studies is through RAAS. RAAS plays an important role in the regulation of blood
pressure and kidney homeostasis. For example, Ang II is an important part of RAAS that stimulates
arteriolar vasoconstriction, leading to increases in blood pressure and hypertension, which as noted above,
could plausibly contribute to kidney dysfunction, CKD, and kidney failure. Following Pb exposure,
vascular reactivity to Ang II was found to increase (Robles et al.. 2007). Exposure to Pb also resulted in
increases in kidney and serum ACE activity as well as renal Ang II-positive cells (Rodriguez-Iturbe et al..
2005; Sharifi et al.. 2004; Carmignani et al.. 1999). Moreover, use of an ACE inhibitor or blocking the
AT-1 receptor (which binds ANG II) ameliorated Pb-induced increases in blood pressure (SimSes et al..
2011).
When considering the available evidence, there are plausible pathways connecting Pb exposure to
renal effects (Figure 5-9). The first potential pathway begins with Pb-induced oxidative stress, which
results in kidney damage and increases in blood pressure, while the second potential pathway is through
the activation of RAAS, which can also result in an increase in blood pressure. Increased blood pressure
can then lead to kidney damage and impaired function, which if sufficiently severe, can lead to kidney
disease. Collectively, these proposed pathways provide biological plausibility for the associations
between Pb levels and adverse renal effects reported in epidemiologic studies.
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5.10
Summary and Causality Determination
In the 2013 Pb ISA, a suggestive relationship between exposure to Pb and reduced kidney
function was judged appropriate on the basis of the health evidence and its associated uncertainties.
Studies published since the 2013 ISA greatly expand the evidence base and serve to strengthen the
evidence for a relationship between exposure to Pb and renal-related health effects. In addition, more
recent evidence has greatly reduced (but not eliminated) key uncertainties from the last review,
particularly those associated with the potential for reverse causality in epidemiologic studies (see below).
This section presents the causality determination for Pb exposures and renal effects, relying upon the
framework for causality determinations described in the Preamble to the ISAs (U.S. EPA, 2015). Key
health evidence supporting this determination is also summarized in Table 5-1.
In the 2013 Pb ISA, prospective epidemiologic studies in older adult, mostly white, men
supported the relationship between long-term Pb exposure and reduced kidney function at mean BLLs
<10 (ig/dL (Tsaih et al., 2004; Kim et al., 1996). Other population-based prospective cohort studies
reported a longitudinal association between BLLs and increases in serum creatinine and CKD progression
over time (Yu et al., 2004). In addition, most epidemiologic cross-sectional studies discussed in the last
review reported that higher tissue Pb concentrations (e.g. blood or bone Pb levels) are associated with
impaired renal function (Navas-Acien et al„ 2009; Muntner et al., 2005; Muntner et al., 2003). Important
uncertainties were raised in the last review with respect to the epidemiologic evidence, particularly the
potential for reverse causality. That is, given the kidney's role in removing toxins from the blood,
increased BLLs could result from reduced excretion due to pre-existing kidney damage rather than as the
causative factor for kidney impairment. It was further noted in the last review that the existence of an
association in adults with normal renal function does not preclude the possibility of reverse causation
because the variation in Pb clearance within the range of normal kidney function is unknown. Other
uncertainties identified in the epidemiologic evidence from the last review were related to the Pb
exposure level, timing, frequency, and duration contributing to the associations reported in these studies
given that most were performed in adult populations with likely higher past Pb exposures. With respect to
the animal toxicology evidence, the 2013 Pb ISA noted that at BLLs >30 (ig/dL, there was clear evidence
that Pb exposure caused changes to the kidney morphology and function (Khalil-Manesh et al„ 1992b;
Khalil-Manesh et al., 1992a). Evidence for functional changes in animals at lower BLLs was more limited
and therefore, more uncertain. When the health evidence was considered along with these uncertainties,
particularly uncertainties related to the potential for reverse causality, the 2013 ISA concluded that
evidence was suggestive of, but not sufficient to infer, a causal relationship between exposure to Pb and
renal effects.
More recent epidemiologic and animal toxicological studies greatly expand the evidence base
from the 2013 Pb ISA. Not only do these newer studies strengthen the evidence of a relationship between
exposure to Pb and renal effects, they also serve to appreciably reduce the uncertainties identified in the
last review. As noted above, the potential for reverse causality was the most influential uncertainty for the
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conclusion in the last review that the scientific evidence was suggestive of, but not sufficient to infer, a
causal relationship between exposure to Pb and renal effects. That is, increased BLLs could result from
reduced excretion due to kidney damage (unrelated to Pb exposure) rather than as a causative factor for
kidney impairment. Cross-sectional studies and studies conducted in populations that are already
experiencing renal dysfunction have the greatest potential for reverse causality. However, prospective
analyses that include both baseline measurements of biomarkers of Pb exposure as well as incident
changes in renal function provide some assurances that associations observed across the epidemiologic
literature are due to a true association with Pb and are not the result of reverse causality. Thus, it is
important to note the more recent longitudinal analyses finding positive associations between exposure to
Pb and kidney disease (Harari et al.. 2018) and decreases in eGFR (Chung et al.. 2020; Liu et al.. 2020).
These longitudinal studies are in agreement with other types of epidemiologic studies reporting similar
associations between exposure to Pb and kidney disease (Wan et al.. 2021; Hagcdoorn et al.. 2020; Lee et
al.. 2020; Wu et al.. 2019; Huang et al.. 2013; Sommar et al.. 2013) and decreases in eGFR (Chung et al..
2020; Liu et al.. 2020; Pollack et al.. 2015). Additional evidence suggesting the that results in
epidemiologic studies are not attributable to reverse causality comes from an epidemiologic study
demonstrating that exposure to Pb is associated with changes in creatinine levels consistent with reduced
kidney function and disease (Pollack et al.. 2015). Importantly, this more recent creatinine study is also
consistent with two longitudinal studies from the prior review presenting similar results (Tsaih et al..
2004; Kim et al.. 1996). These epidemiologic studies were performed in a number of different
geographical areas and in diverse study populations, further reducing the chance that epidemiologic
results are due to reverse causality. Additionally, evidence from RCT trials indicated that an overall
reduction in Pb body burden through EDTA chelation therapy showed evidence of improved renal
function, thus providing more evidence of the effect of Pb on renal outcomes (Lin et al.. 2003; Lin et al..
1999).
Strong support that the associations reported in epidemiologic studies are not from reverse
causality also come from the expanded literature base of animal toxicological studies. In particular, there
is a large body of animal toxicological studies published since the last review largely demonstrating renal
damage or structural abnormalities in rodents following exposure to Pb (Dumkova et al.. 2020a; Gao et
al.. 2020; Shi et al.. 2020; Dumkova et al.. 2017; Alcaraz-Contreras et al.. 2016; Laamech et al.. 2016;
Basgen and Sobin. 2014; Rodriguez-Iturbe et al.. 2005; Fowler et al.. 1980). With respect to
concentrations, effects in rodents were observed in studies at BLLs ranging from -3.0 (ig/dL to
-30 (ig/dL. It is important to note that there is some uncertainty of an effect at this lowest level given that
the same study did not report similar morphological effects at higher BLLs (Basgen and Sobin. 2014) and
that Carlson et al. (2018) reported that renal lesions in mice with a BLL of -3.0 (ig/dL were similar to the
lesions in controls. Nonetheless, there is substantial evidence from animal histological studies for kidney
abnormalities following exposure to Pb, thus providing additional support that the positive associations
for renal disease and impaired renal function reported in longitudinal and cross-sectional epidemiologic
studies are not due to reverse causality. Moreover, these animal toxicology studies also serve to reduce,
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but not eliminate, the uncertainty noted in the last review with respect to effects in animals at the lowest
BLLs.
Epidemiologic studies are also coherent with animal toxicological studies in that they both
provide some evidence of a positive relationship between exposure to Pb and molecular markers of
impaired kidney function in blood, urine, or tissue. As noted above, the 2013 ISA (U.S. EPA. 2013)
evaluated a couple of longitudinal epidemiologic studies that reported positive associations between
increases in serum creatinine levels and bone Pb measurements (Tsaih et al.. 2004; Kim et al.. 1996).
These studies are in agreement with a more recent epidemiologic study describing a positive association
between increasing BLLs and serum creatinine increases in premenopausal women (Pollack et al.. 2015).
In coherence with these epidemiologic studies are a number of animal toxicological studies from the
previous and current review with BLLs below 30 (ig/dL. Although not all studies demonstrated an
increase, most of these studies reported higher blood creatinine levels in Pb-exposed animals compared
with controls (Shi et al.. 2020; Andielkovic et al.. 2019; Laamech et al.. 2016; Zou et al.. 2015; Berrahal
et al.. 2011; Roncal et al.. 2007).
Similar to creatinine levels, changes in measures of blood urea can also be indicative of renal
disease. Although there were no epidemiologic studies examining measures of urea, animal toxicological
studies published since the 2013 Pb ISA (blood Pb values of <30 (ig/dL) generally indicated that exposure
to Pb can increase serum or kidney measures of urea (Gao et al.. 2020; Shi et al.. 2020; Laamech et al..
2016; Zou et al.. 2015). It should be noted, however, that there is at least some variability with respect to
the direction of serum urea levels following Pb exposure. In contrast to the studies mentioned above, both
Andielkovic et al. (2019) and Dumkova et al. (2020a) reported a statistically significant (p <0.05)
decrease in measures of urea relative to controls, while other studies reported no effect (Carlson et al..
2018; Corsetti et al.. 2017) (BLL of 2.89 (ig/dL). It is difficult to interpret whether there is biological
significance to a decrease in serum urea levels relative to control animals, but nonetheless, most animal
toxicological studies reported changes in the levels of urea following exposure to Pb, with most of those
changes being increases. Moreover, the results of these creatinine and urea studies further strengthen the
thesis that the effects observed in epidemiologic studies are truly due to Pb exposure. Other potential
markers of kidney function evaluated in epidemiologic and animal toxicological studies (e.g. UA,
proteinuria) were more limited in number with varying results, and therefore, more uncertain.
As described throughout this causal determination section, there is considerable animal
toxicological evidence supporting Pb as the causative agent for the positive epidemiologic associations
between measures of Pb exposure and adverse health outcomes. Section 5.9 of this document includes
that information to construct a plausible pathway by which exposure to Pb could result in impaired kidney
function or renal disease. In brief, Section 5.10 notes that exposure to Pb can stimulate the production of
reactive oxygen species in the blood or kidneys of exposed laboratory animals (see (U.S. EPA. 2013)
Section 4.5.3.1). Some studies have also reported increases in lipid peroxidation in kidney tissue or
primary cells relative to control animals (Section 5.9). Lipid peroxidation is often an indicator of tissue
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damage and thus, is consistent with the animal histology studies mentioned above demonstrating renal
damage following Pb exposure. Given these results in animal toxicological studies, it is plausible that
associations with renal dysfunction and renal disease (e.g. CKD) reported in epidemiologic studies could
be due to underlying kidney damage from Pb-induced oxidative stress.
The biological plausibility section (Section 5.9) also notes that Pb could potentially lead to the
outcomes reported in epidemiologic studies through RAAS, which has an important role in the regulation
of blood pressure and kidney homeostasis. Ang II is a component of RAAS that stimulates arteriolar
vasoconstriction, leading to increases in blood pressure and hypertension, and Ang II levels can be
increased by exposure to Pb (Sections 5.6 and 5.9). Importantly, prolonged blood pressure increases can
eventually lead to glomerular and renal vasculature injury, plausibly resulting in the renal dysfunction and
renal disease associations observed in epidemiologic studies.
In summary, recent evidence extends the consistency and coherence of the evidence base reported in
the 2013 Pb ISA and is sufficient to conclude that there is a causal relationship between Pb exposure
and renal effects. Recent epidemiologic and animal toxicology studies greatly reduce uncertainties noted
in the previous review, especially with respect to the potential for reverse causality in epidemiologic
studies. Direct evidence for Pb exposure-related renal effects can be found in numerous animal
toxicological studies. In coherence with these results are epidemiologic studies which found that Pb
exposure is associated with some of the same renal endpoints reported in animal toxicological studies
(e.g. eGFR, blood markers of renal impairment). For some markers of renal function, there is a limited
number of studies evaluating these endpoints, and there are some inconsistencies in results across some
of the animal toxicological and epidemiological studies. In general, these studies largely demonstrate a
relationship between exposure to Pb and indicators of kidney distress. Moreover, animal toxicological
studies demonstrating renal damage following Pb exposure provide coherence and biological plausibility
for the consistent epidemiologic associations reported between body Pb concentrations and renal disease.
The key evidence, as it relates to the causal framework, is summarized in Table 5-1.
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Table 5-1 Summary of evidence indicating a causal relationship between Pb
exposure and renal effects
Rationale for
Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels
Associated with Effects0
Generally consistent
Positive associations
(Wu et al.. 2019: Harari et al.. BLLs: -2 to >25
evidence from
between body Pb
2018: Sommar et al.. 2013)
epidemiologic studies
measurements (e.g. blood
of CKD
Pb) and CKD or ESRD
incidence
Generally consistent
evidence from
epidemiologic studies
of diabetic
nephropathy
Mostly positive associations
between body Pb
measurements (e.g. blood
Pb) and diabetic nephropathy
(Wan et al.. 2021: Haaedoorn BLB: <80 to 600 |jg
et al.. 2020: Huang et al..
2013).
BLLs: -1.5 to 6 pg/dL
Generally consistent
Mostly positive associations
(Chung et al.. 2020: Liu et al.. BLLs: -3 to >30 ug/dL
evidence from
between body Pb
2020: Jain. 2019: Buser et al..
epidemiologic studies
measurements (e.g. blood
2016: Chung etal.. 2014: Kim
of eGFR
Pb) and eGFR
and Lee. 2012: Navas-Acien
et al.. 2009: Akesson et al..
2005: Tsaih etal.. 2004: Kim
etal.. 1996)
Generally consistent Mostly positive associations
evidence from between body Pb levels and
epidemiologic studies increases in creatinine
for creatinine in blood
or urine
(Pollack et al.. 2015: Tsaih et
al.. 2004: Kim et al.. 1996)
BLLs: -0.9 to 10 pg/dL
Mostly null findings
from epidemiologic
studies for measures
of UA
Increase in SUA among
women, but not men
Null results between body Pb
and measures of UA and
hyperuricemia
(Park and Kim. 2021)
BLL: -2 pg/dL
(Arrebola et al.. 2019: Jung et BLLs:-0.1 to 2 pg/dL
al.. 2019)
Generally consistent
evidence from animal
toxicological studies
for changes in GFR
Pb-exposed rats had a
statistically significantly lower
(p <0.05) GFR relative to
control rats
(Shi et al.. 2020)
BLL-10.21 pg/dL
Pb-exposed rats had a
statistically significant
increase in GFR indicative of
renal hyperfiltration and
hypertrophy
(Khalil-Manesh etal.. 1993: BLL: -30-45 pg/dL
Khalil-Manesh etal.. 1992b:
Khalil-Manesh etal.. 1992a)
Consistent evidence
Animal toxicological studies
(Gao et al.. 2020: Shi et al.. BLL:~10-30 ug/dL
from animal
consistently demonstrate
2020: Dumkova et al.. 2017:
toxicological studies
renal damage or
Alcaraz-Contreras et al..
of kidney histology
abnormalities in animals
2016: Laamech et al.. 2016:
following Pb exposure
Basgen and Sobin, 2014:
Rodriguez-lturbe et al.. 2005:
Fowler et al.. 1980)
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Rationale for
Causality Key Evidence* References* Associatedwfth
Determination3 Associated witn tnects
Some evidence from
animal toxicological
studies for increased
creatinine in blood or
urine
Most animal studies involving
exposure via drinking water
or gavage demonstrated a
statistically significant
increase in serum creatinine
(or decrease in urine)
following exposure to Pb
A single animal toxicology
study using an inhalation
exposure methodology
reported a decrease in
creatinine levels
(Shi et al.. 2020: Andielkovic BLL:~10-30 |jg/dL
et al., 2019; Laamech et al.,
2016: Zou et al.. 2015)
(Dumkova et al.. 2020a)
BLL: -14 |jg/dL
A couple of animal toxicology g|_|_ 2 89 |jg/dL
studies using a drinking water (Carlson et al.. 2018)
exposure methodology
reported no change in BLL 21.6 pg/dL
creatinine levels in mice (Corsetti etal.,2017)
Some evidence from
animal toxicological
studies for changes
in blood or urine
levels of urea
Most animal studies involving
exposure via drinking water
or gavage demonstrated a
statistically significant
increase in measures of urea
following exposure to Pb
A couple of animal toxicology
studies reported a decrease
in urea levels
An animal toxicology study
reported no change in BUN
levels in mice
(Gao et al.. 2020: Shi et al.. BLL:~10-30 pg/dL
2020: Laamech et al.. 2016:
Zou et al.. 2015)
(Andielkovic et al.. 2019)
(Dumkova et al.. 2020a)
(Carlson et al.. 2018)
BLL:~23 pg/dL
BLL:~14 pg/dL
BLL 2.89 pg/dL
BLB = body lead burden; BLL = blood lead level; BUN = blood urea nitrogen; CKD = chronic kidney disease; eGFR = estimated
glomerular filtration rate; ESRD = end-stage renal disease; GFR = glomerular filtration rate; Pb = lead; SUA = serum uric acid;
UA = uric acid.
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the
Preamble to the ISAs (U.S. EPA. 2015).
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and,
where applicable, to uncertainties or inconsistencies. References to earlier sections indicate where the full body of evidence is
described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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5.11 Evidence Inventories - Data Tables to Summarize Study Details
Table 5-2
Epidemiologic studies of Pb exposure and kidney disease
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs*
Harari et al. (2018)
Malmo,
Sweden
Baseline: 1991-1994,
Follow-up: 2007-2012
Cohort
Cardiovascular Blood Pb (ICP-MS) |jg/dL
cohort of Malmo Diet Median: 2.5 (Range; 0.15-
and Cancer Study 25.8)
(MDCS-CC) Max: 25.8
n = 4,341 enrolled in Quartiles
cohort, 2,567
followed up
Median (range)
Q1 1.5 (0.15-1.85)
Q2 2.2(1.85-2.47)
03 2.9(24.7-3.30)
04 4.6(3.3-25.8)
Q1 +02+ 03 2.2 (0.15—
3.30)
CKD
Age of outcome 73
Linear regression or
Cox proportional
hazards regression
adjusted for age, sex,
smoking, alcohol
intake, hypertension,
diabetes, waist
circumference, eGFR
at baseline, education
level
CKD (HR)a
01 Reference
02 0.83 (0.54, 1.28)
03 0.83 (0.53, 1.29)
04 1.3 (0.85, 2.00)
04 vs. 01 + 02 + 03
1.49 (1.07, 2.08)
Age at measurement 57
Wu et al. (2019)
Taiwan
Case-control
n = 658
220 CKD patients,
438 controls (age
and gender
matched)
Red blood cell Pb
(ICP-MS) (|jg/dL)
Tertiles
T1 <2.794
T2 2.79h4-4.635
T3 >4.635
Age at measurement
Mean (SE)
Cases 65.14 (0.91)
Controls 64.21 (0.60)
CKD
CKD: eGFR
<60 mL/min/1.73 m2 for 3
consecutive mo
Unconditional logistic
regression adjusted
for age, sex,
educational level,
alcohol, tea, and
coffee drinking,
analgesic use,
diabetes,
hypertension, urinary
creatinine, total
urinary arsenic, blood
cadmium, and blood
selenium
Blood Pb log-transformed
ORa
T1 Reference
T2 3.26 (1.58, 6.71)
T3 6.48 (3.23, 12.99)
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs*
Lee et al. (2020)
United States
1999-2016
Cross-sectional (EWAS)
NHANES
n = 46,748
Adults >18
Blood Pb (ICP-MS)
Distribution not reported
Age at Measurement:
Mean (SD) 47 (19)
CKD
CKD 1: ACR >30 mg/g or
eGFR <60 mL/min/1.73 m2
logistic regression
adjusted for age, sex,
diabetes,
hypertension, BMI,
race/ethnicity,
smoking, and SES
CKD 2: ACR >300 mg/g, or
ACR >30 mg/g and eGFR
<60 mL/min/1.73 m2, or eGFR
<45 mL/min/1.73 m2
CKD 3 ACR >300 mg/g and
eGFR <60 mL/min/1.73 m2, or
ACR >30 mg/g and
eGFR <45 mL/min/1.73 m2, or
eGFR <30 mL/min/1.73 m2)
Per SD of the log-
transformed blood Pb
concentration
ORa
CKD 1
Discovery set: 1.27 (1.12,
1.45)
Validation set: 1.12 (1.00,
1.24)
CKD 2
Discovery set: 1.43 (1.29,
1.58)
Validation set: 1.45 (1.29,
1.63)
CKD 3
Discovery set: 1.73 (1.54,
1.95)
Validation set: 1.61 (1.35,
1.90)
Kim et al. (2015)
South Korea
2011
Cross-sectional
KNHANES Blood Pb (GFAAS) (|jg/dL)
n = 1,797 Mean (SD) 2.37 (1.02)
Participants >20 yr of
age
Age at Measurement:
Mean (SD) 46 (14)
CKD (eGFR
<60 mL/min/1.73 m2 or ACR
>30 mg/g)
logistic regression
adjusted for age, sex,
BMI, smoking,
hyperlipidemia,
hypertension,
diabetes, blood
mercury, and blood
cadmium
OR: 1.05 (0.85, 1.30)a
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs*
Sommar et al. (2013)
Sweden
1985 for Vasterbotten
Intervention Project, 1985
for MONICA, 1995 for
Mammography Screening
Project, and 1991-1996 for
Malmo Diet and Cancer
study. Follow-up through
linkage to Swedish Renal
Registry in 2006
Prospective nested case-
referent (mean 7.7 yr of
follow-up, range 1-16 yr)
Vasterbotten
Intervention Project,
the Northern
Sweden WHO
Monitoring of Trends
and Cardiovascular
Disease (MONICA)
study,
Mammography
Screening Project,
and Malmo Diet and
Cancer study
n = 118 cases and
378 controls
Blood (erythrocyte Pb
measured by ICP-MS)
(Hg/dL)
Geometric Mean
Cases 6.62
Referents 5.50
Age at Measurement:
Mean(Range)63 (40-80)
ESRD
(GFR <10-15 mL/min),
starting renal replacement
therapy (i.e., dialysis or
transplantation)
Conditional logistic OR
regression adjusted 1.14(1.03,
for diabetes, BMI, and
hypertension
Three controls
(referents) matched to
each ESRD cases by
cohort, age, sex, and
time of sampling
1.26)
Huang et al. (2013)
China
24-mo observation period
Cohort
n = 85
Patients with type 2
diabetes with
nephropathy (aged
30-83)
eGFR
Primary outcome (2-fold
increase in serum creatinine
from baseline values, need for
long-term dialysis, or death)
BLB (X-ray fluorescence and Diabetic Nephropathy
EDTA) (pg)
Low (BLB <80 |jg)
Mean (SD) 58.1 (16.7)
Max: 79.8
High (BLB 80-600 pg)
Mean (SD) 132.4 (46.1)
Max 316.8
Blood (ETAAS) (|jg/dL)
Low (BLB <80 |jg)
Mean (SD) 3.8 (3.0)
Max 10.4
High (BLB 80-600 pg)
Mean (SD) 4.6 (3.1)
Max: 10.3
Longitudinal
eGFR (mL/min/1.73 m2
1 pg increase in BLB
-0.022 (-0.039, -0.005)
1 pg/dL increase in Blood
Pb
-0.298 (-0.525, -0.071)
multivariate analysis
or Cox regression
analysis adjusting for
age, sex, smoking,
BMI, history of CVD,
MAP, cholesterol,
triglycerides, HbA1c,
serum creatinine, daily Primary outcome
protein intake, daily BLB: HR: 1.01 (95% CI:
protein excretion 1.01,1.02)
BLB >80 pg: HR2.79 (CI
1.25, 6.25)
Age at Measurement
Mean (SD) 60.1 (9.5) Range
33-83
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs*
Haqedoorn et al. (2020)
The Netherlands
2009-2016
Cross-sectional
Blood Pb (ICP-MS) (pg/dL) DKD
DIAbetes and
LifEstyle Cohort
Twente-1 (DIALECT- (|QR) ^ (Q ^
n = 231 1-86>
With type 2 diabetes Age at Measurement:
Mean (SD) 64 (9)
(Creatinine clearance
<60 mL/min/1.73 m2) and/or
presence of albuminuria (ALB
excretion >30 mg/d)
Logistic regression OR for doubling of blood
adjusted for age, sex, Pb (log 2 transformed)
HbA1c, insulin use, yr (pmol/L)a
of diabetes, MAP,
alcohol intake, pack Creatinine clearance
yr, and blood <60 mL/min/1.73 m2
cadmium OR 1.83 (1.07, 3.15)
Albuminuria > 30 mg/d
OR 1.75 (1.11, 2.74)
Wan et al. (2021)
China
May-August 2018
Cross-sectional
Environmental
Pollutant Exposure
and Metabolic
Diseases in
Shanghai n = 4,234
Blood (Atomic Absorption
Spectrometry) Pb (pg/dL)
Median (IQR)
2.6 (1.8, 3.6)
Age at Measurement
Median (IQR)
67 (62-72) yr
DKD Linear or logistic
regression adjusting
ACR (high, >30 mg/g); DKD as for age, sex, duration
defined by American Diabetes of diabetes, education
Association (ACR >30 mg/g or status, current
eGFR <60 mL/min per smoking, BMI, HbA1c,
1.73 m2) dyslipidemia,
hypertension
OR (4th vs. 1st quartile of
Blood Pb)a
DKD 1.36 (1.06, 1.74)
ACR (>30 mg/g) 1.31,
(1.02, 1.69))
Hara et al. (2016)
Northeastern Belgium
Baseline blood Pb (1985-
1989), follow-up through
2014
Cohort
Cadmium in Belgium Blood Pb (ETAAS with
(CadmiBel) study
n = 1,302
Flemish residents
(>20 yr), randomly
recruited from 10
districts in
northeastern
Belgium
Zeeman correction) (pg/dL)
Geometric Mean (IQR) 6.00
(3.31, 10.35)
Age at Measurement:
Mean (SD) 47.8 (15.6)
Nephrolithiasis (Self-reported Cox regression
and verified by investigators adjusted for age, sex,
against medical records. serum magnesium,
Cases were symptomatic, and 24 hr urinary volume,
often hospitalized for diagnosis and calcium
and treatment)
Per doubling of blood Pb
(pmol/L)a(HR)
Baseline Pb 1.35 (1.06,
1.73)
Mean (baseline and follow-
up averaged): 1.32 (1.03,
1.71)
Baseline with regression
dilution bias correction
1.44 (1.07, 1.93)
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% CIs*
Sun etal. (2019)
NHANES
Blood Pb (ICP-MS)d (|jg/dL)
Nephrolithiasis (Self-reported
Logistic regression
Weighted OR (95% CI)
n = 21,402
Median: 1.22
history of kidney stones)
adjusting for age, sex,
2007-2016
95th: 3.89
race/ethnicity, BMI,
Compared with reference
Adult (>20 yr)
Max: 61.29
educational level,
level (0.05 pg/dL)
Cross-sectional
participants from
marital status, annual
0.50 pg/dL: 0.88 (0.81,
NHANES
family income,
0.95)
smoking, physical
1.00 pg/dL: 0.75 (0.63,
activity, intake of total
0.89)
energy, calcium,
1.50 pg/dL: 0.67 (0.52,
phosphate, sodium,
0.85)
potassium,
2.00 pg/dL: 0.62 (0.46,
magnesium, total fluid,
0.83)
alcohol, caffeine,
2.5 pg/dL: 0.60 (0.44,
vitamin B6, vitamin C,
0.82)
and vitamin D, and
3.0 pg/dL: 0.60 (0.43,
eGFR
0.84)
3.5 pg/dL: 0.60 (0.44,
0.86)
4.0 pg/dL: 0.61 (0.44,
0.86)
4.5 pg/dL: 0.63 (0.45,
0.88)
5.0 pg/dL: 0.64 (0.46,
0.90)
ACR = albumin-to-creatinine ratio; ALB = albumin; BLB = body lead burden; BMI = body mass index; CI = confidence interval; CKD = chronic kidney disease; CVD = cardiovascular
disease; DKD = diabetic kidney disease; eGFR = estimated glomerular filtration rate; EDTA = ethylenediaminetetraacetic acid; ESRD = end-stage renal disease;
ETAAS = Electrothermal Atomic Absorption Spectrometry; EWAS = environment wide association study; GFAAS = graphite furnace atomic absorption spectrometry;
GFR = glomerular filtration rate; HbA1c = hemoglobin A1c; HR = hazard ratio; hr = hour(s); ICP-MS = inductively coupled plasma mass spectrometry; IQR = interquartile range;
MAP = mean arterial pressure; MDCS-CC = cardiovascular cohort of the Malmo Diet and Cancer Study; mo = month(s); MONICA = Monitory of Trends and Cardiovascular Disease;
NHANES = National Health and Nutrition Examination Survey; OR = odds ratio; Pb = lead; Q = quartile; SE = standard error; SES = socioeconomic status; T# = fertile #; yr = year(s).
'Effect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect estimates
are standardized to the specified unit increase for the 10th-90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval.
Categorical effect estimates are not standardized.
aUnable to be standardized.
increment unclear.
Confidence intervals estimated based on reported p-values.
dBlood Pb analysis method not reported, assumed based on data set (NHANES).
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Table 5-3
Animal toxicological studies of Pb exposure and kidney histology
Study
Species (Stock/Strain), n, Timing of
Sex Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (ng/dl_)
Endpoints Examined
Basaen and
Sobin (2014)
Mouse
Control
(Pb-free drinking water),
M/F, n = 12, (6/6)
30 ppm, M/F, n = 12, (6/6)
330 ppm, M/F, n = 12, (6/6)
In utero to Drinking water from
PND 28 dams was treated with
99.4% Pb acetate. Litters
were then exposed to 0,
30, or 330 Pb acetate in
drinking water for 28 d
0.03 ± 0.01 |jg/dL for control
males
0.03 ± 0.01 |jg/dL for control
females
3.63 |jg/dL± 0.71 pg/dLfor
30 ppm males
2.74 pg/dL ± 0.36 pg/dL for
30 ppm females
16.02 pg/dL± 3.25 pg/dLfor
330 ppm males
Kidney Histology, podocyte
characteristics and glomerular
volume post 4-wk exposure
13.35 pg/dL ± 1.31 pg/dLfor
330 ppm females
Li etal. (2017)
Mouse (Balb/c)
Control
(water), F, n = 8
100 mg/kg/d Pb acetate, F,
n = 8
6-7 wk old mice
8 wk
Plain water or
100 mg/kg/d Pb acetate
for 1 d then given skim
milk from d 2-15
0.43 ± 0.05 pg/L for control
(4.3 ± 0.05 pg/dL)
302.20 ± 25.32 pg/L for
100 mg/kg/d Pb acetate
(30.2 ± 25.32 pg/dL)
Kidney Histology post
exposure
Alcaraz-
Contreras et al.
(2016)
Rat (Wistar)
Control (water), M, n = 5
2,000 ppm Pb acetate, M,
n = 5
2 mo old rats
exposed to Pb
for 8 wk
2 mo old rats received
drinking water, or
drinking water with
2000 ppm Pb acetate for
8 wk
21.9 ±2.0 pg/dLfor
2000 ppm group
Kidney Histologyl d post I
wk exposure
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Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (ng/dl_)
Endpoints Examined
Rahman et al.
(2018)
Rat (Wistar)
Control (tap water), M/F,
n = 7-8
0.2% Pb acetate, M/F,
n = 7-8/group
PND 1 to
PND 30
Pups were exposed to
0.2% Pb acetate from
PND 1 to PND 21
through dam's drinking
water. Then rats were
exposed directly through
drinking water until
PND 30. Control animals
were given tap water
throughout
2.2 ± 0.7 |jg/dL for control-
PND 21
12.4 ± 3.3 |jg/dL for 0.2% Pb
acetate-PND 21
3.3 ± 1.7 |jg/dL for control-
PND 30
22.7 ± 6.0 |jg/dL for 0.2% Pb
acetate-PND 30
Kidney Histology at PND 21
and PND 30
Andielkovic et
al. (2019)
Rat (Wistar)
Control water, M, n = 8
150 mg/kg b.w., M, n = 6
Single exposure
by oral gavage
(age of rats not
reported)
Single oral dose of
150 mg/kg b.w. Pb
acetate
-25 |jg/L for control
(-2.5 pg/dL)
-225 pg/L for 150 mg/kg b.w.
Pb acetate
(-22.5 pg/dL)
Kidney histology 24-hr post
exposure
Carlson et al.
(2018)
Mouse (Control)
(water), M/F, n = 16
0.03 mM Pb, M/F, n = 8
Treatment began
no earlier than
an age
of 5 wk for 11 wk
Pb-free water or
0.03 mM Pb acetate
dissolved in drinking
water for 11 wk
Control (water) not detected
2.89 ± 0.44 pg/dLfor
0.03 mM
Kidney Histology one wk after
11 wk exposure
Dumkova et al.
(2017)
Mouse
Experiment 1:
Control (clean air), F,
n = 5
Adult mice
exposed for 6 wk
Experiment 1:
1.23 x 10s particles/cm3
of PbO inhalation
exposure or clean air for
6 wk (24/hr a day, 7 d
a week)
Experiment 2:
<11 ng/g for control
(<1.166 pg/dL)
132 ng/g for Pb-exposed
(13.992 pg/dL; not specified
from which experiment
measurement was derived)
Kidney Histology post 6 wk
exposure
5-49
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (ng/dl_)
Endpoints Examined
1.23 x 10® PbO
particles/cm3, F, n = 5
Experiment 2:
Control (clean air), F,
n = 5
0.956 x 10s particles/cm3,
F, n = 5
0.956 x 10s particles/cm3
of PbO inhalation
exposure or clean air for
6 wk (24/hr d, 7 d a wk)
(Experiment 2 was a
replicant of experiment
1):
Laamech et al.
(2016)
Mouse
Control
(distilled water), M/F,
n = 10
5 mg/kg/d Pb acetate,
M/F, n = 10
Age of mice in
experiment not
reported
Distilled water or
5 mg/kg/d Pb acetate
dissolved in distilled
water for 40 d
0.009 |jg/ml_ for control
(distilled water) (0.9 pg/dL)
0.18 |jg/ml_ for 5 mg/kg/d Pb
acetate (18 pg/dL)
Kidney Histology 2 d post
exposure
Shi et al.
(2020)
Rat (SD)
Control (deionized water),
M, n = 8
0.5% Pb acetate, M, n = 8
28 d after
PND21
After 21 d of milk
feeding, 0.5% Pb acetate
or deionized water for
28 d
0.18 ± 0.07 |jg/dL for Control
(deionized water)
10.21 ± 0.93 |jg/dL for 0.5%
Pb acetate
Kidney Histology post
exposure
Gao et al.
(2020)
Rat (SD)
Control
(Distilled water), M/F,
n = 10
5 mg/kg Pb acetate, M/F,
n = 10
Age of mice in
experiment not
reported
5 mg/kg Pb acetate orally
for 35 d followed by
recovery to d 63
<0.02 mg/kg for distilled water
(<2.12 |jg/dL)
0.10 ± 0.03 mg/kg for 5 mg/kg
Pb acetate (d 64)
(10.6 ± 0.03 |jg/dL)
Kidney Histology following the
end of the experiment on d 63
5-50
-------
Study
Species (Stock/Strain), n, Timing of
Sex Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (ng/dl_)
Endpoints Examined
Dumkova et al. Mouse (Control)
(clean air), F, n = 10 (wk,
6 wk, 11 wk)
PbO, F, n = 10 (2 wk, 6 wk,
11 wk)
PbO recovery, F, n = 10
(6 wk PbO, 5 wk clean air)
Age of mice in
experiment
unclear
PbO 78.0 |jg PbO/m3 or
clean air for 24 hr/d
7 d/wk for 2 wk, 6 wk, or
11 wk. A recovery group
was exposed to PbO for
6 wk and then clean air
for 5 wk (11 wk total)
<3 ng/g in control (2 wk, 6 wk,
11 wk) (0.3 |jg/dL)
104 ng/g PbO 2 wk
(10.4 |jg/dL)
148 ng/g PbO 6 wk
(14.8 |jg/dL)
174 ng/g PbO 11 wk
(17.4 Mg/dL)
Kidney histology at 2 wk,
6 wk, and 11 wk
Dumkova et al.
(2020a)
Mouse (Control)
(clean air), F, n = 10 (d 3,
2 wk, 6 wk, 11 wk)
Pb(N03)2 (68.6 |ag/m3), F,
n = 10 (d 3, 2 wk, 6 wk,
11 wk)
Recovery (Pb(N03)2
68.6 |ag/m3), F, n = 10
(6 wk Pb/5 wk recovery)
6-8 wk old mice
exposed for 3 d,
2 wk, 6 wk, or
11 wk
Pb(NOs) (68.6 |ag/mA3)
or clean air-exposed
mice for 3 d, 2 wk, 6 wk,
or 11 wk. To assess
recovery, a separate
group of mice were
exposed for 11 wk
followed by 5 wk of clean
air
<0.3 ng/g for control at all
timepoints (<0.3 |ag/dL)
(d 3, 2 wk, 6 wk, 11 wk)
31 ng/g for Pb(N03)2 d 3
(3.1 ng/dL)
40 ng/g for Pb(N03)2 2 wk
(4.0 |ag/dL)
47 ng/g for Pb(N03)2 6 wk
(4.7 |ag/dL)
Kidney Histology post 3 d,
2 wk, 6 wk, 11 wk, and 11 wk
plus clearance for 5 wk
(-16 wk)
8 5 ng/g for Pb(N03)2 11 wk
(8.5 |ag/dL)
10 ng/g forPb(N03)2
exposure 6 wk and clean air
for 5 wk (1.0 |jg/dL)
d = day(s); hr = hour(s); mo = month(s); M = male; M/F = male/female; N03 = nitrate, PND = postnatal day, Pb(N03)2 = Pb nitrate, PbO = Pb oxide; wk = week(s).
5-51
-------
Table 5-4
Epidemiologic studies of Pb exposure and estimated glomerular filtration rate
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs*
Yu et al. (2004)
Adult CKD
Blood Pb (ETAAS with
Change in eGFR
Cox proportional hazard model
Change in eGFR per 1 |jg/dL
patients
Zeeman correction) (pg/dL)
(MDRD) over 4 yr
examined whether a predictor
increase in blood Pb
Taipei, Taiwan; 48-mo
longitudinal study
period
n = 121
Mean (SD)
4.2 (2.2)
(mL/min/1.73 m2)
was associated with renal
function including age, sex,
BMI, hyperlipidemia,
-4.01 (-7.24, —0.78)a
10th—90th percentile
2.0-5.1
hypertension, smoking, use of
ACE inhibitor, baseline serum
Cohort
creatinine, daily protein
excretion daily protein intake,
underlying kidney disease
Harari et al. (2018)
Malmo,
Sweden
Baseline: 1991-1994,
Follow-up: 2007-2012
Cohort
Cardiovascular
cohort of Malmo
Diet and Cancer
Study (MDCS-
CC)
n = 4,341
enrolled in
cohort, 2,567
followed up
Blood Pb (ICP-MS) |jg/dL
Median: 2.5 (Range; 0.15-
25.8)
Max: 25.8
Quartiles
Median (range)
Q1 1.5 (0.15-1.85)
Q2 2.2 (1.85-2.47)
Q3 2.9 (24.7-3.30)
Q4 4.6 (3.3-25.8)
Q1 +Q2 + Q3 2.2 (0.15—
3.30)
Change in eGFR
(CKD-EPI) from
baseline
Age at outcome 73
Linear regression adjusted for
age, sex, smoking, alcohol
intake, hypertension, diabetes,
waist circumference, eGFR at
baseline, education level
Change in eGFRc
(mL/min/1.73m2)
Q1 (Reference)
Q2 -1.70 (-3.10, -0.26)
Q3 -2.90 (-4.30, -1.50)
Q4 -2.30 (-3.80, -0.73)
Age of measurement 57
5-52
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs*
Liu et al. (2020)
Shiyan City of Hubei
Province
China
Baseline between
September-June
2010, follow-up in
2013
Mean follow-up: 4.6 yr
Cohort
Dongfeng-
Tongji
n = 1,434
Retirees from
Blood Pb (ICP-MS) |jg/dL
Median (IQR)
1.23 (0.84-1.90)b
Quartiles
Q1 <0.843
Dongfeng Motor Q2 0.843-1.232
Corporation
Q3 1.232-1.895
Q4 >1.895
Age at Measurement:
Mean (SD) 62.4 (7.5)
Annual eGFR
(CKD-EPI) decline
([Baseline eGFR-
eGFR at follow-
up]/number of follow-
up years)
Linear regression adjusted for
age, sex, baseline eGFR,
batch (from the 3 case-
controls), occupational
category, BMI, smoking status,
drinking status, education
level, and fasting plasma
glucose
Annual decline in eGFR
(mL/min/1.73 m2) per In-
transformed increase in blood
Pbcd
Q1 Referent
Q2 0.30 (-0.20, 0.81)
Q3 0.30 (-0.20, 0.81)
Q4 0.83 (0.31, 1.35)
Chung et al. (2020) n = 770
Taiwan
Recruited 2010-2011
and follow-up in 2015-
2016
Cohort
Community
residents living
near an EAF
Blood Pb (ICP-MS) (pg/dL)
Geometric mean (IQR)
Distance from EAF
<500 m 2.41 (1.22-6.19)
500-1000 m 2.26 (1.16-
4.83)
1000-1500 m 2.12 (1.05-
4.67)
1500-2000 m 2.23 (0.98—
4.31)
>2000 m: 2.03 (1.03-4.31)
eGFR (method not
specified)
General linear models
adjusting for age, sex,
ethnicity, living near the main
road and smoking
Per 1 pg increase in blood Pb:
(mL/min/1.73 m2)
eGFR: -2.25 (-3.50, -1.01)
Age at measurement
Median 60
5-53
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs*
Pollack et al. (2015)
Buffalo, NY
United States
women followed
2 menstrual cycles (8 f°r 2 menstrual
visits per cycle) 2005- cyc'es
2007
Cohort
BioCycle Blood Pb (ICP-MS) (pg/dL) eGFR (MDRD)
n = 259 Median (IQR) 0.86 (0.68—
1.2)
Premenopausal Mean (SD) 1 03 (0.63)
Age at Measurement:
Mean (SD) 27.4 (8.2)
Linear mixed models adjusted
forage, BMI, race, average
calories, alcohol intake,
smoking, and cycle day
Percent Change in Kidney
Biomarkers per 2-fold increase
in blood Pbd
eGFR: -3.73 (-6.55, -0.83)
OR
eGFR (<90 mL/min/1.73 m2)
0.32 (0.08, 1.21)
eGFR (<60 mL/min/1.73 m2)
0.32 (0.08, 1.21)
Results presented as percent
change in nontransformed
outcome per 2-fold increase in
nontransformed exposure
Navas-Acien et al.
(2009)
United States
1999-2006
Cross-sectional
NHANES
adults
n = 14,778
Aged >20 yr
Blood Pb (ICP-MS) (pg/dL)
Geometric mean 1.58
Quartiles
Range (Median)
Q1
Q2
Q3
Q4
<1.1 (0.8)
1.2 to 1.6 (1.3)
1.7 to 2.4 (1.9)
>2.4 (3.2)
Reduced eGFR
(MDRD) (eGFR
<60 mL/min/1.73 m2
Logistic regression adjusted for OR
survey year, age, sex,
race/ethnicity, BMI, education,
smoking, cotinine, alcohol
intake, hypertension, diabetes,
menopausal status
Q1 Referent
Q2 1.21 (0.64, 2.28)
Q3 1.32 (1.00, 1.76)
Q4 1.20 (1.07, 1.36)
Age of measurement
Mean (SD)
Reduced eGFR 67.6 (0.5)
No reduced eGFR 44.7 (0.3)
5-54
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs*
Muiai etal. (2019)
United States
May 2015 to
September 2017
Cross-sectional
SPHERL
n = 447 men
Newly hired Pb
workers at
battery
manufacturing
and Pb
recycling plants
Blood Pb (ICP-MS) (|jg/dL)
Geometric mean (IQR)
1.66 (1.3-2.5)
Age at Measurement:
Mean (SD) 28.7 (10.2)
eGFR (CKD-EPI),
ACR
Linear regression adjusted for
age, MAP, BMI, smoking,
waist-to-hip ratio, total
cholesterol to HDL ratio,
plasma glucose, y-glutamyl
transferase, and
antihypertensive drug
treatment
Per doubling of blood Pb
(mL/min/1.73 m2)d
eGFRcrt (serum creatinine)
-0.135 (-3.40, 3.13)
eGFRcys (serum cystatin)
-0.222 (-3.07, 2.62)
eGFRcc (serum creatinine and
cystatin): -0.281 (-3.07, 2.50)
Per doubling of blood Pb
(mg/mmol)
ACR: -0.071 (-0.14, 0.59)
Kim and Lee (2012)
South Korea
2008-2010
Cross-sectional
KNHANES
n = 5,924
Participants
>20 yr of age
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
Geometric mean (95% CI)
2.289 (95% CI: 2.258,
2.319)
Quartiles
Q1 <1.743
Q2 >1.734-2.305
Q3 >2.305-3.010
Q4 >3.010
eGFR (MDRD)
(Considered reduced
if <80 mL/min per
1.73 m2)
Linear and logistic regression
adjusted for age, sex,
residence area, education
level, smoking status, drinking
status, hypertension, diabetes,
hemoglobin, blood cadmium,
and blood mercury
Continuous eGFRd
(mL/min/1.73 m2)
Doubling of Pb
-2.624 (-3.803
1.445)
Q1 Reference
Q2 -0.491 (-2.048, 1.0651)
Q3 -2.341 (-4.013, -0.669)
Q4 -3.835 (-5.730, -1.939)
Reduced eGFR (OR (95% Cl))d
Doubling of Pb
1.324 (1.139, 1.540)
Q1 Reference
Q2 1.031 (0.806, 1.319)
Q3 1.161 (0.892, 1.511)
Q4 1.631 (1.246, 2.136)
5-55
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs*
Chung et al. (2014)
South Korea
2008
Cross-sectional
KNHANES
n =2,005
>20 yr with data
for blood Pb and
cadmium.
Pregnant
women were
excluded
Blood Pb (GFAAS with
Zeeman correction) (pg/dL)
Geometric mean: 2.5
Quartiles (Mean)
Q1 1.38
Q2 2.10
Q3 2.74
Q4 4.13
Age at Measurement:
Mean(Range)46(20-87)
eGFR (CKD-EPI)
Linear regression adjusted for
age, sex, smoking,
hypertension, or diabetes.
Logistic regression adjusted for
age, sex, smoking
hypertension, BMI, and blood
cadmium
Per 1 |jg/dL increase in blood
Pb (mL/min/1.73 m2)
-2.61 (95% CI: -3.29, -1.97)
OR (95% CI) (Q4 vs. Q1, per
1 |jg/dL increase in blood Pb
eGFR (<60 mL/min/1.73 m2)
1.08 (95% CI: 0.99, 1.17)
Buser et al. (2016)
United States
2007-2012
Cross-sectional
NHANES
n =4,875
Pregnant and
breastfeeding
women were
excluded
Blood Pb (ICP-MS) (pg/dL)
Quartiles
Q1 <0.79
Q2 0.80-1.20
Q3 1.21-1.82
Q4 >1.82 pg/dL
Age at Measurement
Geometric Mean 44.1
eGFR (CKD-EPI)
Linear regression adjusting for
age, race/ethnicity, sex,
diabetes, alcohol intake,
education, smoking status,
body weight, hypertension,
weak/failing kidney, serum
cotinine, and blood cadmium
eGFR (mL/min/1.73 m2)d
Q1 Reference
Q2 -1.17 (-2.91, 0.57)
Q3 -1.62 (-3.60, 0.36)
Q4 -2.67 (-4.78, -0.56)
Jain (2019) NHANES Blood Pb (ICP-MS) (pg/dL) Reduced eGFR Logistic regression adjusting OR (95% Cl)cd
n = 25,427 (CKD-EPI) for sex, race/ethnicity, smoking
United States 75th percentile: 2.15 (<60 mL/min/1.73 m2) status, age, BMI, survey year, eGFR (<60 mL/min/1.73 m2):
>20 yr of age fasting time, poverty income 1.567 (1.346, 1.823)
9nm-9ni4 „ x ™ ratio, diabetes, and
Age at measurement >20 yr hypertension
Cross-sectional
5-56
-------
Refe,enD?SfgnnS'Udy Population Exposure Assessme„,
Outcome
Confounders
Effect Estimates and 95% CIs*
Per SD of the log-transformed
blood Pb concentration11
OR eGFR (<60 mL/min/1.73 m2)
Discovery set: 1.35 (1.24, 1.48)
Validation set 1.27 (1.11, 1.45)
OR eGFR (<45 mL/min/1.73 m2)
Discovery set: 1.60 (1.39, 1.85)
Validation set 1.63 (1.42, 1.88)
OR eGFR (<30 mL/min/1.73 m2)
Discovery set: 1.98 (1.50, 2.62)
Validation set 2.25 (1.75, 2.90)
ACE = angiotensin-converting enzyme ACR = albumin-to-creatinine ratio; BMI = body mass index; CI = confidence interval; CKD = chronic kidney disease; CKD-EPI = Chronic
Kidney Disease Epidemiology Collaboration; EAF = electric arc furnace; eGFR = estimated glomerular filtration rate; ETAAS = Electrothermal Atomic Absorption Spectrometry;
GFAAS = graphite furnace atomic absorption spectrometry; HDL = high-density lipoprotein; ICP-MS = inductively coupled plasma mass spectrometry; IQR = interquartile range;
KNHANES = National Health and Nutrition Examination Survey; MAP = mean arterial pressure; MDCS-CC = cardiovascular cohort of the Malmo Diet and Cancer Study;
MDRD = Method of eGFR calculation from the Modification of Diet in Kidney Disease study; mo = month(s); NHANES = National Health and Nutrition Examination Survey;
OR = odds ratio; Q = quartile; SD = standard deviation; SES = socioeconomic status; SPHERL = Study for Promotion of Health in Recycling Lead; Pb = lead; yr = year(s).
'Effect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
Confidence interval estimated from reported p-value.
bUnits converted from |jg/L.
°lncrement unclear.
dUnable to be standardized.
Lee et al. (2020)
United States
1999-2016
Cross-sectional
NHANES
n = 46,748
Adults >18 yr of
age
Blood Pb (ICP-MS)
Distribution not reported
Age at Measurement:
Mean (SD) 47 (19)
Reduced eGFR
(CKD-EPI) (<60, <45,
or
<30 mL/min/1.73 m2)
Logistic regression adjusted for
age, sex, diabetes,
hypertension, BMI,
race/ethnicity, smoking, and
5-57
-------
Table 5-5
Animal toxicological studies of Pb exposure and glomerular filtration rate
Study
Species
(Stock/Strain),
n, Sex
Timing of Exposure
Exposure
Details
(Concentration,
Duration)
BLL as Reported (iig/dL
Endpoints Examined
Shi et al. (2020)
Rat (SD)
Control
(deionized
water), M, n = 8
0.5% Pb acetate,
M, n = 8
28 d after PND 21
After21 d of 0.18 ± 0.07 |jg/dL for Control GFR postexposure
milk feeding, (deionized water)
0.5% Pb acetate
or deionized 10 21 ± 0.93 pg/dL for 0.5% Pb
water for 28 d acetate
d = day(s); GFR = glomerular filtration rate; M = male; Pb = lead; PND = postnatal day; SD = standard deviation.
5-58
-------
Table 5-6
Epidemiologic studies of Pb exposure and albumin, creatinine, and albumin-to-creatinine ratio
Reference and Study Study
Design Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs*
Tsaih et al. (2004)
Boston, MA
1991-1995, ~4 yr of
follow-up
Cohort
NAS
n = 448
Adult males,
mostly white
Blood Pb (Blood (GFAAS
with Zeeman correction)
(pg/dL)
Mean (SD) 6.5 (4.2)
10th—90th 2.1-7.6
Bone Pb (K-XRF) (pg/g)
Mean (SD)
Tibia 21.5 (13.5)
Patella 32.4 (20.5)
Age at measurement
Mean (SD) 66.0 (6.6)
Change in
serum creatinine
(mg/dL) peryr
Age at outcome
Mean (SD) 72.0
(6.5)
Log linear regression
adjusted for age, BMI,
hypertension, diabetes,
smoking status, alcohol
consumption, analgesic use,
baseline serum creatinine
Annual change in serum creatinine
(mg/dL/yr)
Blood Pb
Overall 0.002 (-0.001, 0.004)
Diabetic 0.013 (0.005, 0.02)
Nondiabetic 0.001 (0, 0.002)
Hypertensive 0.001 (-0.002, 0.005)
Normotensive 0.002 (0, 0.003)
Tibia Pb
Overall, 0.035 (-0.014, 0.084)
Diabetic 0.412 (0.146, 0.678)
Nondiabetic 0.025 (-0.024, 0.074)
Hypertensive 0.116 (0.017, 0.214)
Normotensive 0.002 (-0.057, 0.061)
Kim et al. (1996)
Boston, MA
1979-1994
Retrospective cohort
NAS
n = 459
Adult males,
mostly white
Blood Pb (Blood (GFAAS
with Zeeman correction)
(pg/dL)
Median 8.6
10th—90th percentile: 4.0-
17.5
Change in Random-effects modeling
Serum adjusted for baseline age,
creatinine time since initial visit, BMI,
(mg/dL) smoking status, alcohol
ingestion, education level,
and hypertension
Change in serum creatinine (mg/dL)
Peak blood Pb <40 pg/dL
0.0017 (0.0005, 0.003)
Peak blood Pb <25 pg/dL
0.0021 (0.0007, 0.0035)
Peak blood Pb <10 pg/dL
0.0033 (0.0012, 0.0053)
Akesson et al. (2005) WHILA, adult
women
Sweden n = ®^0
1999-2000
Cross-sectional
Median (5%-95% CI)
concurrent blood Pb: 2.2
(1.1, 4.6) pg/dL
10th—90th percentile: 1.3-3.1
Creatinine
clearance/100
(mL/min)
Linear regression adjusted
for age, BMI, diabetes,
hypertension, regular use of
nephrotoxic drug, smoking
status
Creatinine clearance/100 (mL/min)
for each unit increase in blood Pb
-0.018 (-0.03, -0.006)
5-59
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs*
Pollack et al. (2015)
Buffalo, NY
United States
2 menstrual cycles (8
visits per cycle) 2005-
2007
Cohort
BioCycle
n = 259
Premenopausal
women followed
for 2 menstrual
cycles
Blood Pb (ICP-MS) (pg/dL)
Median (IQR) 0.86 (0.68-
Mean (SD) 1.03 (0.63)
Age at Measurement:
Mean (SD) 27.4 (8.2)
1.2)
Creatinine and
ALB
(BUN, CO2,
Chloride,
Potassium,
Urate, Calcium,
Protein,
Glucose)
Linear mixed models
adjusted forage, BMI, race,
average calories, alcohol
intake, smoking, and cycle d
Percent Change in kidney
Biomarkers per 2-fold increase in
blood Pba
Creatinine: 3.47 (0.85, 6.16)
ALB -0.38 (-1.28, 0.52)
BUN: -0.13 (-4.97, 4.96)
C02: -0.57 (-1.43, 0.29)
Chloride: 0.20 (-0.09, 0.48)
Potassium: 0.01 (-1.15, 1.18)
Urate: 0.90 (-2.22, 4.12)
Calcium: -0.21 (-0.67, 0.25)
Protein: -0.76 (-1.61, 0.09)
Glucose: 0.93 (-0.28, 2.15)
*Results presented as percent
change in nontransformed outcome
per 2-fold increase in
nontransformed exposure
Buser et al. (2016)
United States
2007-2012
Cross-sectional
NHANES
n = 4,875
Pregnant and
breastfeeding
women were
excluded
Blood Pb (ICP-MS) (pg/dL)
Quartiles
Q1 <0.79
Q2 0.80-1.20
Q3 1.21-1.82
Q4 >1.82 pg/dL
Age at Measurement:
Geometric Mean 44.1
Urinary ALB Linear regression adjusting
for age, race/ethnicity, sex,
diabetes, alcohol intake,
education, smoking status,
body weight, hypertension,
weak/failing kidney, serum
cotinine, and blood cadmium
ALB (percent difference)3
Q1 Reference
Q2 -4.02 (-13.76, 6.93)
Q3 -9.24 (-19.43, 2.22)
Q4 6.29 (-6.39, 20.80)
5-60
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs*
Muiai etal. (2019)
SPHERL
Blood Pb (ICP-MS) (pg/dL) ACR
Linear regression adjusted
Per doubling of blood Pb
n = 447 men
Geometric mean (IQR)
forage, MAP, BMI, smoking,
(mg/mmol)a
United States
1.66 (1.3-2.5)
waist-to-hip ratio, total
ACR:
Newly hired Pb
cholesterol to HDL ratio,
-0.071 (-0.14, 0.59)
May 2015 to
September 2017
workers at
Age at Measurement:
plasma glucose, y-glutamyl
battery
transferase, and
manufacturing
Mean (SD) 28.7 (10.2)
antihypertensive drug
Cross-sectional
and Pb recycling
plants
treatment
Jain (2019)
United States
2003-2014
Cross-sectional
NHANES
n = 25,427
>20 yr
Blood Pb (ICP-MS) (pg/dL) ACR
75th percentile 2.15
Age at measurement >20 yr
Logistic regression adjusting
for sex, race/ethnicity,
smoking status, age, BMI,
survey year, fasting time,
poverty income ratio,
diabetes, and hypertension
OR (95% Cl)ab
ACR (>30 mg/g creatinine)
1.206 (1.05, 1.385)
Zhu etal. (2019)
United States
2009-2012
Cross-sectional
NHANES
n = 2926
>20 yr
Blood Pb (ICP-MS) (pg/dL)
Quartiles
Q1 <0.0685
Q2 0.0686-0.1029
Q3 0.1030-0.1600
Q4 >0.1601
Age at Measurement:
Mean (SE)42.1 (0.46)
ACR
Linear regression adjusted Blood Pb and continuous ACR (In-
forage, sex, BMI, obesity,
ethnicity, education,
smoking, hypertension,
diabetes, and CKD
transformed) (mg/g)a
Q1 Reference
Q2 0.04 (-0.06, 0.13)
Q3 -0.05 (-0.18, 0.08)
Q4 0.06 (-0.08, 0.20)
5-61
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs*
Lee et al. (2020)
United States
1999-2016
Cross-sectional
NHANES Blood Pb (ICP-MS)
n = 46,748 Distribution not reported
Adults >18 Age at Measurement:
Mean (SD) 47 (19)
ACR (>30 and Logistic regression adjusted Per SD of the log-transformed blood
>300 mg/g) for age, sex, diabetes,
hypertension, BMI,
race/ethnicity, smoking, and
Pb concentration3
OR ACR (>30 mg/g)
Discovery set: 1.23(1.07, 1.42)
Validation set 1.08 (0.97, 1.20)
OR ACR (>300 mg/g)
Discovery set: 1.39 (1.22, 1.59)
Validation set 1.38 (1.16, 1.63)
ACR = albumin-to-creatinine ratio; ALB = albumin; BMI = body mass index; BUN = blood urea nitrogen; CI = confidence interval; CKD = chronic kidney disease; GFAAS = graphite
furnace atomic absorption spectrometry; ICP-MS = inductively coupled plasma mass spectrometry; IQR = interquartile range; K-XRF = K-shell X-ray Fluorescence; MAP = mean
arterial pressure; NHANES = National Health and Nutrition Examination Survey; OR = odds ratio; Pb = lead; Q = quartile; SD = standard deviation; SES = socioeconomic status;
SPHERL = Study for Promotion of Health in Recycling Lead; yr = year(s).
'Effect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
aUnable to be standardized.
blncrement unclear.
5-62
-------
Table 5-7
Animal toxicological studies of Pb exposure and albumin and creatinine
Study
Species (Stock/Strain), Timing of
n, Sex Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (ng/dl_)
Endpoints Examined
Zou etal. (2015)
Mouse (Control)
(re-distilled water), M,
n = 10
3-wk exposure
of
approximately
30-d-old
250 mg/L Pb acetate, M, mice
n = 10
250 mg/L Pb acetate 1.8 pg/dL for Control (re-distilled water)
or distilled water for
3 wk
21.7 |jg/dL for 250 mg/L-PND 58
Markers of Kidney Function:
Creatinine post 3-wk
exposure
Corsetti et al. (2017) Mouse (Control) d 30 to d 75
(Pb-free water), M, n = 8
200 ppm Pb, M, n = 8
Mice were exposed to <5 Mg/dL for 0 ppm
ordinary or Pb
21.6 pg/dL for 200 ppm
Markers of Kidney Function:
serum creatinine post 45-d
exposure
Andielkovic et al.
(2019)
Rat (Wistar)
Control water, M, n = 8
150 mg/kg b.w., M, n = 6
Single
exposure by
oral gavage
(age of rats not
reported)
Single oral dose of
150 mg/kg b.w. Pb
acetate
-25 |jg/L for Control (-2.5 pg/dL)
-225 pg/L for 150 mg/kg b.w. Pb
acetate
(-22.5 Mg/dL)
Markers of Kidney Function:
serum levels of creatinine 24
hr post single exposure
Zinc and copper levels in the
kidney 24 hr post single
exposure
Shi et al. (2020) Rat (SD) 28 d after
Control (deionized water), 21
M, n = 8
0.5% Pb acetate, M,
After 21 d of milk
feeding, 0.5% Pb
acetate or deionized
water for 28 d
0.18 ± 0.07 pg/dL for Control (deionized GFR and Markers of Kidney
water) Function: Creatinine post
exposure
10.21 ± 0.93 pg/dL for 0.5% Pb acetate
5-63
-------
Study
Species (Stock/Strain), Timing of
n, Sex Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (ng/dl_)
Endpoints Examined
Laamech et al.
(2016)
Mouse
Control
(distilled water), M/F,
n = 10
5 mg/kg/d Pb acetate,
M/F, n = 10
Age of mice in
experiment not
reported
Distilled water or
5 mg/kg/d Pb acetate
dissolved in distilled
water for 40 d
0.009 |jg/ml_ for control (distilled water)
(0.9 pg/dL)
0.18 pg/mL for 5 mg/kg/d Pb acetate
(18 pg/dL)
Markers of Kidney Function:
plasma levels of creatinine
2 d post exposure
Gao et al. (2020) Rat (SD)
Age of mice in 5 mg/kg Pb acetate <0.02 mg/kg for distilled water
Control
(Distilled water), M/F,
n = 10
5 mg/kg Pb acetate, M/F,
n = 10
experiment not orally for 35 d
reported followed by recovery
to d 63
(<2.12 pg/dL)
0.10 ± 0.03 mg/kg for 5 mg/kg Pb
acetate (d 64) (10.6 ± 0.03 pg/dL)
Markers of Kidney Function:
creatinine activity following
the end of the experiment on
d 63
Dumkova et al.
(2020b)
Mouse (Control)
(clean air), F, n = 10
(2 wk, 6 wk, 11 wk)
PbO, F, n = 10 (2 wk,
6 wk, 11 wk)
PbO recovery, F, n = 10
(6 wk PbO, 5 wk clean
air)
Age of mice in PbO 78.0 pg PbO/m3 <3 ng/g in control (2 wk, 6 wk, 11 wk)
experiment or clean air for 24 hr/d (0.3 pg/dL)
unclear 7 d/wk for 2 wk, 6 wk,
or 11 wk. a recovery
group was exposed to
PbO for 6 wk and
then clean air for 5 wk 148 ng/g PbO 6 wk (14.8 pg/dL)
(11 wk total)
174 ng/g PbO 11 wk (17.4 pg/dL)
104 ng/g PbO 2 wk (10.4 pg/dL)
Markers of Kidney Function:
Creatinine at 2 wk, 6 wk, and
11 wk
d = day(s); GFR = glomerular filtration rate; hr = hour(s); F = female; M = male; M/F = male/female; Pb = lead; PbO = Pb oxide; PND = postnatal day; SD = standard deviation;
wk = week(s).
5-64
-------
Table 5-8
Epidemiologic studies of Pb exposure and uric acid3
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
Park and Kim (2021)
South Korea
2016-2017
Cross-sectional
KNHANES
n =4,784
Participants
>20
Blood Pb (GFAAS) (pg/dL)
Geometric mean:
Overall, 1.69
Men 1.95
Women 1.50
Age at measurement >20
SUA and Linear and logistic regression
hyperuricemia adjusting for age, residence
(SUA area, education level, smoking
>7.0 mg/dL in status, drinking status, physical
men and activity, hypertension, glucose,
>6.0 mg/dL in triglyceride, cholesterol, eGFR,
women) blood cadmium and blood
mercury
Per doubling of Blood Pb
Log SUA (mg/dL)
Men: -0.018 (-0.038, 0.002)
Women: 0.019 (0.001, 0.037)
Hyperuricemia (OR) per doubling
of blood Pba
Men: 0.928 (0.718, 1.198)
Women: 1.095 (0.727, 1.649)
Arrebola et al. (2019)
BIOAMBIENT.E
Blood Pb (method not
UA and
logistic or linear regression
Per 1 unit increase in log-
S study
indicated) (pg/dL)
hyperuricemia
adjusting for sex, age, weight
transformed Pb
Spain
n = 882
(UA >7.0 mg/dL
loss in past 6 mo, smoking
2009-2010
Cross-sectional
458 males and
424 females
Median 0.106
75th 0.181
in males or
>6.0 mg/dL in
status, alcohol consumption,
education, region of
Log SUA (mg/dL)
90th 0.284
females,
recruitment, place of residence
5.95 (-0.02, 0.05)
95th 0.355
prescribed any
medication for
Hyperuricemia (OR)
Age at measurement
lowering UA
levels, diagnosis
1.12 (0.90, 1.41)
Median 35.4-38.1
of gout by a
physician)
5-65
-------
Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
Jung etal. (2019)
South Korea
2016
KNHANES
n = 2,682
1124 men and
1528 women)
aged >19 yr
Blood Pb (GFAAS) (pg/dL)
Hyperuricemia
Median (IQR) 2.04 (1.59-2.51)
No Hyperuricemia
Median (IQR) 1.73 (1.34-2.28)
Age at measurement
Hyperuricemia
Mean (SE) 46.4 (1.3)
No hyperuricemia
Mean (SD) 46.9 (0.5)
Hyperuricemia Logistic regression adjusting for See Figure 5-5
(SUA
>7.0 mg/dL in
men and
>6.0 mg/dL in
women)
age, BMI, eGFR, residence,
education, smoking status,
alcohol consumption, physical
activity, and blood pressure
BMI = body mass index; CI = confidence interval; eGFR = estimated glomerular filtration rate; IQR = interquartile range; GFAAS = graphite furnace atomic absorption spectrometry;
KNHANES = Korea National Health and Nutrition Examination Survey; mo = month(s); OR = odds ratio; SD = standard deviation; SE = standard error; SUA = serum uric acid;
UA = uric acid; yr = year(s).
aUnable to be standardized.
5-66
-------
Table 5-9 Animal toxicological studies of Pb exposure and measures of uric acid and urea
Study
Species
(Stock/Strain),
n, Sex
Timing of Exposure
Exposure
Details
(Concentration,
Duration)
BLL as Reported (ng/dl_)
Endpoints Examined
Zou etal. (2015)
Mouse (Control) 3-wk exposure of
(re-distilled approximately 30-d-old
water), M, n = 10 mice
250 mg/L Pb
acetate, M,
n = 10
250 mg/L Pb
acetate or
distilled water
for 3 wk
1.8 |jg/dL for Control (re-
distilled water)
21.7 ug/dL for 250 mg/L-
PND 58
Markers of Kidney Function: BUN
post 3-wk exposure
Andielkovic et al.
(2019)
Rat (Wistar)
Control water,
M,
150 mg/kg b.w.,
M, n = 6
Single exposure by oral
gavage (age of rats not
reported)
Single oral dose
of 150 mg/kg
b.w. Pb acetate
-25 |jg/L for Control
(-2.5 pg/dL)
-225 pg/L for 150 mg/kg b.w.
Pb acetate
(-22.5 pg/dL)
Markers of Kidney Function:
serum levels of BUN 24 hr post
single exposure
Zinc and copper levels in the
kidney 24 hr post single
exposure
Shi et al. (2020)
Rat (SD)
Control
(deionized
water), M, n =
0.5% Pb
acetate, M, n :
28 d after PND 21
After 21 d of
milk feeding,
0.5% Pb acetate
or deionized
water for 28 d
0.18 ±0.07 pg/dL for Control
(deionized water)
10.21 ± 0.93 pg/dL for 0.5% Pb
acetate
Markers of Kidney Function: BUN
and UA post exposure
Carlson et al. (2018)
Mouse (Control)
(water), M/F,
n = 16
0.03 mM Pb,
M/F, n = 8
Treatment began no earlier
than an age
of 5 wk for 11 wk
Pb free water or Control (water) not detected
0.03 mM Pb
acetate
dissolved in
drinking water
for 11 wk
2.89 ± 0.44 pg/dL for 0.03 mM
Markers of Kidney Function: BUN
1 wk after 11-wk exposure
5-67
-------
Study
Species
(Stock/Strain),
n, Sex
Timing of Exposure
Exposure
Details
(Concentration,
Duration)
BLL as Reported (ng/dl_)
Endpoints Examined
Laamech et al. (2016) Mouse
Control
Age of mice in experiment
not reported
(distilled water),
M/F, n = 10
5 mg/kg/d Pb
acetate, M/F,
n = 10
Distilled water or
5 mg/kg/d Pb
acetate
dissolved in
distilled water
for 40 d
0.009 |jg/ml_ for control (distilled
water) (0.9 pg/dL)
0.18 |jg/ml_ for 5 mg/kg/d Pb
acetate (18 pg/dl_)
Markers of Kidney Function:
plasma levels of urea and UA 2 d
post exposure
Gao et al. (2020)
Rat (SD)
Control
(Distilled water),
M/F,
n = 10
5 mg/kg Pb
acetate, M/F,
n = 10
Age of mice in experiment
not reported
5 mg/kg Pb
acetate orally for
35 d followed by
recovery to d 63
<0.02 mg/kg for distilled water
(<2.12 |jg/dL)
0.10 ± 0.03 mg/kg for 5 mg/kg
Pb acetate (d 64)
(10.6 ± 0.03 |jg/dL)
Markers of Kidney Function: BUN
activity following the end of the
experiment on d 63
Dumkova et al. (2020b)
Mouse (Control)
(clean air), F,
n = 10 (2 wk,
6 wk, 11 wk)
PbO, F, n = 10
(2 wk, 6 wk,
11 wk)
PbO recovery,
F, n = 10 (6 wk
PbO, 5 wk clean
air)
Age of mice in experiment
unclear
PbO 78.0 |jg
PbO/m3 or clean
air for 24 hr/d
7 d/wk for 2 wk,
6 wk, or 11 wk.
a recovery
group was
exposed to PbO
for 6 wk and
then clean air for
5 wk (11 wk
total)
<3 ng/g in control (2 wk, 6 wk,
11 wk) (0.3 |jg/dL)
104 ng/g PbO 2 wk
(10.4 |jg/dL)
148 ng/g PbO 6 wk
(14.8 |jg/dL)
174 ng/g PbO 11wk
(17.4 pg/dL)
Markers of Kidney Function:
Urea at 2, 6, and 11 wk
BUN = blood urea nitrogen; d = day(s); F = female; hr = hour(s); M = male; M/F = male/female; Pb = lead; PbO = Pb oxide; SD = standard deviation; wk = week(s).
5-68
-------
Table 5-10 Epidemiologic studies of Pb exposure and proteinuria and hematuria
Reference and Study Study
Design Population
Exposure Assessment Outcome
Confounders
Effect Estimates and 95% CIs*
Chung et al. (2014)
South Korea
2008
Cross-sectional
KNHANES
n = 2,005
>20 yr with data
for blood Pb and
cadmium.
Pregnant women
were excluded
Blood (GFAAS with Zeeman
correction) (pg/dL)
Geometric mean: 2.5
Quartiles (Mean)
Q1 1.38
Q2 2.10
Q3 2.74
Q4 4.13
Age at Measurement:
Mean(Range)46 (20-87)
Proteinuria Linear regression adjusted for
age, sex, smoking,
hypertension, or diabetes.
Logistic regression adjusted for
age, sex, smoking
hypertension, BMI, and blood
cadmium
OR (95% CI) (Q4 vs. Q1, per
1 |jg/dL increase in blood Pb)
1.08 (1.00, 1.16)
Han et al. (2013)
South Korea
2008-2010
Cross-sectional
KNHANES Blood (GFAAS with Zeeman
n = 4,701 correction) (pg/dL)
Geometric mean: 1.08
Quartiles
Q1 <1.89
Q2 1.89-2.46
Q3 2.47-3.22
Q4 >3.22
Age at Measurement:
Mean 50 yr
Hematuria Logistic regression adjusting for
age, sex, residential region,
education level, and anemia
• OR
(95% Cl)a
Q1
Reference
Q2
1.00 (0.767,
1.303)
Q3
0.90 (0.687,
1.178)
Q4
0.94 (0.701,
1.253)
BMI = body mass index; BUN = blood urea nitrogen; CI = confidence interval; GFAAS = graphite furnace atomic absorption spectrometry; KNHANES = Korea National Health and
Nutrition Examination Survey; OR = odds ratio; Q = quartile; yr = year(s).
'Effect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
aUnable to be standardized.
5-69
-------
Table 5-11
Epidemiologic studies of Pb exposure and renal tubular impairment markers3
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs
~ metal. (2016)
KRIEFS
Blood Pb (GFAAS)
Renal tubular
Linear regression adjusting for
Log-transformed Pb
n = 1953
(pg/dL)
impairment
age, sex, BMI, household
South Korea
Geometric mean 2.21
(NAG and p2-
income, smoking, alcohol
NAG (Unit/g creatinine)
2010-2012
participants >19
MG)
consumption, hypertension, and
0.09 (-0.05, 0.23)
without kidney
Age at Measurement
diabetes
Cross-sectional
disease
Mean 45.5
p2-MG (pg/g creatinine)
0.01 (-0.13, 0.15)
Jung etal. (2016)
Jangseong-gun
South Korea
June-August 2013 and
August-November
2014
Cross-sectional
n = 547
Participants living
near cement plant
Blood Pb (Atomic
Absorption Spectrometry
[flameless method])
(pg/dL)
Quartiles
Q1 0.77-2.13
Q2 2.13-2.70
Q3 2.70-3.50
Q4 3.50-10.37
Renal tubular
impairment
(NAG
>5.67 U/L)
Logistic regression adjusting for Renal tubular impairment
sex, age, occupation, smoking,
air pollution exposure,
hypertension, diabetes, urine
cadmium, urine mercury
OR (95% CI)
Q1 Reference
Q2 0.96 (0.49,
Q3 0.89 (0.44,
Q4 0.67 (0.32,
1.87)
1.77)
1.41)
Age at Measurement:
Mean (SD) 64.32 (11.02)
(B2-MG = (32-microglobulin; BMI = body mass index; CI = confidence interval; GFAAS = graphite furnace atomic absorption spectrometry; KRIEFS = Korean Research Project on the
Integrated Exposure Assessment to Hazardous Materials for Food Safety; NAG = acetyl-(B-D-glucosaminidase; OR = odds ratio; Q = quartile; SD = standard deviation.
aUnable to be standardized.
5-70
-------
Table 5-12
Animal toxicological studies of Pb exposure and other markers of kidney function
Study
Species
(Stock/Strain),
n, Sex
Timing of Exposure
Exposure
Details
(Concentration,
Duration)
BLL as Reported (ng/dL>
Endpoints Examined
Andielkovic et al.
Rat (Wistar)
Single exposure by oral
Single oral dose
-25 |jg/L for Control
Total protein, zinc, and copper
(2019)
Control water,
gavage (age of rats not
of 150 mg/kg
(-2.5 pg/dL)
levels in the kidney 24 hr post
M, n = 8
reported)
b.w. Pb acetate
single exposure
-225 pg/L for 150 mg/kg b.w.
150 mg/kg b.w.,
Pb acetate
M, n = 6
(-22.5 pg/dL)
Fioresi et al. (2014)
Rat (Wistar)
Control (tap
water), M,
n = 9-12
Age 2 mo to 3 mo
100 ppm Pb
acetate in
drinking water
for 30 d
<0.5 pg/dL for control
13.6 ± 1.07 pg/dL for
100 ppm group
ACE activity measured post 30-d
exposure
100 ppm group,
M,
n = 9-12
5-71
-------
Exposure
Species Details
Study (Stock/Strain), Timing of Exposure ,_ . .. BLL as Reported (ng/dU Endpoints Examined
n sex (Concentration,
' Duration)
Dumkova et al. (2020a) Mouse
6-8 wk old mice exposed for Pb (N03)2
(Control)
(clean air), F,
n = 10 (d 3,
2 wk, 6 wk,
11 wk)
Pb(N03)2
(68.6 ug/mA3),
F, n = 10 (d 3,
2 wk, 6 wk,
11 wk)
Recovery
(Pb(N03)2
68.6 pg/m3), F,
n = 10 (6 wk
Pb/5 wk
recovery)
3 d, 2 wk, 6 wk, or 11 wk
or
(68.6 pg/m3
clean air-
exposed mice
for 3 d, 2 wk,
6 wk, or 11 wk.
To assess
recovery a
separate group
of mice were
exposed for
11 wk followed
by 5 wk of clean
air
<0.3 ng/g for control at all
timepoints (<0.3 pg/dL)
(d 3, 2 wk, 6 wk, 11 wk)
31 ng/g for Pb(N03)2 d 3
(3.1 pg/dL)
40 ng/g for Pb(N03)2 2 wk
(4.0 pg/dL)
47 ng/g for Pb(N03)2 6 wk
(4.7 pg/dL)
85 ng/g for Pb(N03)2 11 wk
(8.5 pg/dL)
10 ng/g for Pb(N03)2 exposure
6 wk and clean air for 5 wk
(1.0 pg/dL)
Total protein, calcium, sodium,
and potassium levels in the
kidney post 3 d, 2 wk, 6 wk,
11 wk, and 11 wk plus clearance
for 5 wk (-16 wk)
5-72
-------
Study
Species
(Stock/Strain),
n, Sex
Timing of Exposure
Exposure
Details
(Concentration,
Duration)
BLL as Reported (ng/dL>
Endpoints Examined
Dumkova et al. (2020b)
Mouse
(Control)
(clean air), F,
n = 10 (2 wk,
6 wk, 11 wk)
PbO, F, n = 10
(2 wk, 6 wk,
11 wk)
PbO recovery,
F, n = 10 (6 wk
PbO, 5 wk
clean air)
Age of mice in experiment
unclear
PbO 78.0 |jg
PbO/m3 or clean
air for 24 hr/d
7 d/wk for 2 wk,
6 wk, or 11 wk.
A recovery
group was
exposed to PbO
for 6 wk and
then clean air for
5 wk (11 wk
total)
<3 ng/g in control (2 wk, 6 wk,
11 wk) (0.3 |jg/dL)
104 ng/g PbO 2 wk
(10.4 |jg/dL)
148 ng/g PbO 6 wk
(14.8 |jg/dL)
174 ng/g PbO 11 wk
(17.4 |jg/dL)
Total protein post 2 wk, 6 wk,
and 11 wk exposure
d = day(s); hr = hour(s); F = female; M = male; PbO = Pb oxide; wk = week(s).
5-73
-------
Table 5-13 Epidemiologic studies of Pb exposure and renal outcomes in children
Reference and Study
Design
Study
Population
Exposure Assessment Outcome
Confounders
Effect Estimates and 95% CIs*
Fadrowski et al. (2010) NHANES
n = 769
1988-1994
Cross-sectional
Adolescents
aged 12-20
Whole blood Pb
(GFAAS) (pg/g)
Median (IQR) 1.5 (0.7,
2.9)
Quartile
Q1 <1.0
Q2 1.0-1.5
Q3 1.6-2.9
Q4 >2.9
Age at measurement
12-15 46%
16-20 54%
eGFR (cystatin C
and serum
creatinine-based
estimates)
Linear regression
adjusted for age, sex,
race/ethnicity, urban vs.
rural residence, tobacco
smoke exposure,
obesity, annual
household income, and
educational level of the
family reference person
eGFR (mL/min/1.73 m2)
referent (Q1 )a
Cystatin C-based
Q2 -1.4 (-7.4, 4.5)
Q3 -2.6 (-7.3, 2.2)
Q4 -6.6 (-12.6, -0.7)
Creatinine-based
Q2 -0.5 (-6.1, 5.1)
Q3 -1.7 (-6.9, 3.5)
Q4 -1.9 (-7.4, 3.5)
compared with
Skroder et al. (2016)
Bangladesh
June 2002-June 2004
Cohort
Maternal and
Infant Nutrition
Interventions,
Matlab
n = 1,511
(GW14); 713
(GW30)
Mother-child
pairs
Erythrocyte Pb (ICP-MS
followed by dilution in
alkali solution (GW14)
or acid digestion
(GW30)) (pg/g)
GW14
Median (95th) 73 (172)
GW30
Median (95th) 86 (506)
Age at Measurement:
Mean (SD) 26 (6) (age
of mothers)
Kidney volume,
eGFR, serum
cystatin C
Blood pressure in
children
Age at outcome
4.5 yr
Linear regression
adjusted for gender,
birth weight, season of
birth, age at outcome
measurements, height
for age Z-score,
maternal BMI at GW8,
parity, SES, and
supplementation
group
Per pg/kg Eyr-Pba
GW14
Kidney volume (cm3/m2) 0.062 (-0.36, 0.24)
eGFR (mL/min/1.73 m2) 0.089 (-0.012,0.30)
Serum cystatin C (mg/L)
-0.00088 (-0.0028, 0.001)
GW30
Kidney Volume (cm3/m2)
-0.071 (-1.4, -0.030)
eGFR (mL/min/1.73 m2) 0.71 (-0.24,0.17)
Serum cystatin C (mg/L)
0.000027 (-0.0018, 0.0018)
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Reference and Study
Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95% CIs*
Fadrowski et al. (2013)
CkiD
Whole blood Pb (high
GFR
Linear regression
Change (mL/min/1.73 m2) in GFR
n = 391 (485 Pb
resolution ICP-MS)
adjusted for age, sex,
-0.9 (-2.6, 0.8)
United States and
measurements)
(pg/dL)
GFR directly
race, ethnicity, BMI,
Canada
Median (Range) 1.2
(0.2-6.2)
measured at yr 2
poverty, CKD diagnosis
Percent change in GFR
Children with
and 4 of CkiD
(glomerular or
2007-2009
CKD (1-16 yr of
age)
study
nonglomerular), urine
protein to creatinine
Total sample -2.1 (-6.0, 1.8)
With glomerular CKD -12.1 (-22.2, -1.9)
Cross-sectional
Age at measurement
0-5 13%
6-11 38%
12-19 49%
ratio, and In-
transformed blood
cadmium
With nonglomerular CKD -0.7 (-4.8, 3.4)
Cardenas-Gonzalez et
n = 83
Whole blood Pb
Kidney Injury
Linear regression
No associations between blood Pb and
al. (2016)
Children
(GFAAS) (pg/dL)
Molecule 1
(KIM-1) and
adjusted for age, sex,
BMI, urinary specific
kidney injury biomarkers (data not shown)
San Luis Potosi
attending 2
Geometric mean
neutrophil
gravity, or urinary
Mexico
elementary
(Range)
gelatinase-
creatinine
schools in San
5.95 (1.47-26.89)
associated
2014
Luis Potosi,
Mexico
lipocalin (NGAL)
Age at Measurement
Cross-sectional
Mean (SD) 8.13 (1.93)
5-75
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Reference and Study
Design
Study
Population
Exposure Assessment Outcome
Confounders
Effect Estimates and 95% CIs*
Hu et al. (2019)
NHANES
Blood Pb (Atomic SUA
Linear regression
Per 1 unit increase in In-transformed blood
n = 8,303
Absorption
adjusted age, sex, BMI,
Pba
United States
Spectrometry with
race, education status,
SUA (mg/dL) 0.14 (0.10, 0.17)
Adolescents
Zeeman correction)
hemoglobin, HDL-C,
1999-2006
aged 12-19
(pg/dL)
and eGFR
OR (SUA >5.5 mg/dL) 1.29 (1.17, 1.42)
Mean: 1.3
Cross-sectional
Age at Measurement
Mean (SD) 15.5 (2.3)
BMI = body mass index; CKD = chronic kidney disease; CKiD = Chronic Kidney Disease in Children; eGFR = estimated glomerular filtration rate; GFAAS = graphite furnace atomic
absorption spectrometry; GFR = glomerular filtration rate; GW = gestational week; HDL-C = high-density lipoprotein cholesterol; ICP-MS = inductively coupled plasma mass
spectrometry; IQR = interquartile range; KIM-1 = kidney injury molecule 1; NGAL = neutrophil gelatinase-associated lipocalin; NHANES = National Health and Nutrition Examination
Survey; OR = odds ratio; Pb = lead; Q = quartile; SD = standard deviation; SES = socioeconomic situation; SUA = serum uric acid; UA = uric acid; yr = year(s).
'Effect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th-90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
aUnable to be standardized.
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United States
Environmental Protection
Agency
Integrated Science
Assessment for Lead
Appendix 6: Immune System Effects
January 2024
EPA/600/R-23/375
January 2024
www.epa.gov/isa
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
-------
CONTENTS
DOCUMENT GUIDE 6-iii
LIST OF TABLES 6-v
LIST OF FIGURES 6-vi
ACRONYMS AND ABBREVIATIONS 6-vii
APPENDIX 6 IMMUNE SYSTEM EFFECTS 6-1
6.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current Review 6-1
6.2 Scope 6-3
6.3 Immunosuppression 6-5
6.3.1 Epidemiologic Studies of Immunosuppression 6-5
6.3.2 Toxicological Studies of Immunosuppression 6-9
6.3.3 Integrated Summary of Immunosuppression 6-18
6.4 Sensitization and Allergic Responses 6-20
6.4.1 Epidemiologic Studies of Sensitization and Allergic Responses 6-21
6.4.2 Toxicological Studies of Sensitization and Allergic Responses 6-23
6.4.3 Integrated Summary of Sensitization and Allergic Responses 6-24
6.5 Autoimmunity and Autoimmune Disease 6-25
6.5.1 Epidemiologic Studies of Autoimmunity and Autoimmune Disease 6-25
6.5.2 Toxicological Studies of Autoimmunity and Autoimmune Disease 6-26
6.5.3 Integrated Summary of Autoimmunity and Autoimmune Disease 6-26
6.6 Biological Plausibility 6-27
6.6.1 Immunosuppression 6-29
6.6.2 Sensitization and Allergic Responses 6-31
6.7 Summary and Causality Determinations 6-32
6.7.1 Causality Determination for Immunosuppression 6-32
6.7.2 Causality Determination for Sensitization and Allergic Responses 6-37
6.7.3 Causality Determination for Autoimmunity and Autoimmune Disease 6-40
6.8 Evidence Inventories - Data Tables to Summarize Study Details 6-42
6.9 References 6-76
6-iv
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LIST OF TABLES
Table 6-1 Summary of evidence for a likely to be causal relationship between Pb exposure and
immunosuppression 6-35
Table 6-2 Summary of evidence that is suggestive of, but not sufficient to infer, a causal relationship
between Pb exposure and sensitization and allergic responses 6-39
Table 6-3 Summary of evidence that is inadequate to determine the presence or absence of a
causal relationship between Pb exposure and autoimmunity and autoimmune disease 6-41
Table 6-4 Epidemiologic studies of exposure to Pb and immunosuppression 6-42
Table 6-5 Animal toxicological studies of delayed-type hypersensitivity responses 6-51
Table 6-6 Animal toxicological studies of antibody response 6-52
Table 6-7 Animal toxicological studies of ex vivo white blood cell function 6-52
Table 6-8 Animal toxicological studies of immune organ pathology 6-54
Table 6-9 Animal toxicological studies of immunoglobulin levels 6-56
Table 6-10 Animal toxicological studies of immune organ weight 6-57
Table 6-11 Animal toxicological studies of white blood cell counts and differentials (spleen, thymus,
lymph node, bone marrow) 6-63
Table 6-12 Animal toxicological studies of white blood cell counts (hematology and subpopulations) 6-65
Table 6-13 Epidemiologic studies of exposure to Pb and sensitization and allergic response 6-66
Table 6-14 Animal toxicological studies of immediate-type hypersensitivity 6-73
Table 6-15 Epidemiologic studies of exposure to Pb and autoimmunity and autoimmune disease 6-74
Table 6-16 Animal toxicological studies of autoimmunity and autoimmune disease 6-75
6-v
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LIST OF FIGURES
Figure 6-1 Potential biological plausibility pathways for immunological effects associated with
exposure to Pb. 6-28
6-vi
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ACRONYMS AND ABBREVIATIONS
AQCD
anti-TT
BLL
BMI
BW
Cd
CD
CI
CMI
Con A
CR1
d
DNFB
DTH
e-waste
EDEN
EE
EGFP
ELISA
F
Fe
GFAAS
GM-CSF
hr
HBc
HBsAb
HBV
Hib
HLA-DR
HR
ICR
ICP-MS
IFN-y
Ig-
IL-
ILC
ILCP
ISA
ISO
KNHANES
Air Quality Criteria Document
anti-tetanus toxoid
blood lead level
body mass index
body weight
cadmium
cluster of differentiation
confidence interval
cell-mediated immune
Concanavalin A
complement receptor type 1
day(s)
1 -Fluoro-2,4-dinitrobenzene
delayed-type hypersensitivity
electronic-waste
Effect of Diet and Exercise on
Immunotherapy and the Microbiome
effect estimate
enhanced green fluorescent protein
enzyme-linked immunosorbent assay
female
iron
graphite furnace atomic absorption
spectrometry
granulocyte-macrophage colony-
stimulating factor
hour, hours
hepatitis B core
hepatitis B surface antigen
hepatitis B virus
Haemophilus influenzae type B
Major histocompatibility complex
(MHC) II cell surface receptor
hazard ratio
Institute for Cancer Research
inductively coupled plasma mass
spectrometry
interferon-gamma
immunoglobulin type -
interleukin type -
innate lymphoid cell
innate lymphoid cell progenitor
Integrated Science Assessment
isolation
Korea National Health and Nutrition
Examination Survey
In
M
MMR
M/F
min
mo
MRSA
MSSA
NHANES
NK
NO
NR
OR
Pb
PbONP
PCR
PECOS
PND
ppm
Q
ROS
RR
RSV
S/CO
SCORAD
SD
SES
SPT
STELLAR
T#
TDAR
Th2
TNF
Treg
TSLP
IT
tTG
WBC
wk
yr
vs.
natural log
male
measles, mumps, and rubella
male/female
minute(s)
month(s)
methicillin-resistant Staphylococcus
aureus
methicillin-sensitive Staphylococcus
aureus
National Health and Nutrition
Examination Survey
natural killer
nitric oxygen
not reported
odds ratio
lead
lead oxide nanoparticle
polymerase chain reaction
Population, Exposure, Comparison,
Outcome, and Study Design
postnatal day
parts per million
quartile
reactive oxygen species
relative risk
respiratory syncytial virus
signal to cut-off
scoring atopic dermatitis
standard deviation
socioeconomic status
skin prick test
Systemic Tracking of Elevated Lead
Levels and Remediation
tertile #
T cell dependent antibody response
T cell-derived helper cell 2
tumor necrosis factor
regulatory T cell
thymic stromal lymphopoietin
tetanus toxoid
tissue transglutaminase
white blood cell
week(s)
year(s)
versus
6-vii
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APPENDIX 6 IMMUNE SYSTEM EFFECTS
Causality Determinations for Pb Exposure and Immune System Effects
This appendix characterizes the scientific evidence that supports causality
determinations for lead (Pb) exposure and immune system effects. The types of studies
evaluated within this appendix are consistent with the overall scope of the ISA as
detailed in the Process Appendix (see Section 12.4). In assessing the overall evidence,
the strengths and limitations of individual studies were evaluated based on scientific
considerations detailed in Table 12-5 of the Process Appendix (Section 12.6.1). More
details on the causal framework used to reach these conclusions are included in the
Preamble to the ISA (U.S. EPA. 2015). The evidence presented throughout this
appendix supports the following causality conclusions:
Outcome Group Causality Determination
Immunosuppression Likely to be Causal
Sensitization and Allergic Suggestive of, but not sufficient to infer, a
Responses causal relationship
Autoimmunity and , , .
... J. Inadequate
Autoimmune Disease M
The Executive Summary, Integrated Synthesis, and all other appendices of this Pb ISA
can be found at https://assessments.epa.gov/isa/document/&deid=359536.
6.1 Introduction, Summary of the 2013 Pb ISA, and Scope of
the Current Review
The 2013 Integrated Science Assessment for Lead (hereinafter referred to as the 2013 Pb ISA)
issued causality determinations for the effects of Pb exposure on different aspects of the immune system
including atopic and inflammatory responses, decreased host resistance, and autoimmunity (U.S. EPA.
2013). It is not without precedent for a single chemical to exert both stimulatory and suppressive effects
on various immune parameters (IPCS. 2012). The evidence underpinning these causality determinations is
briefly summarized below.
The body of epidemiologic and toxicological evidence integrated across the 2013 Pb ISA
indicates a "likely to be causal" relationship between Pb exposure and increased atopic and inflammatory
conditions. This relationship is supported by evidence for associations of blood Pb levels (BLL) with
asthma and allergy in children and Pb-associated increases in immunoglobulin E (IgE) in children and
laboratory animals. Uncertainties in the epidemiologic evidence related to potential confounding by
6-1
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socioeconomic status (SES), smoking, or allergen exposure are reduced through consideration of the
evidence from experimental animal studies. The biological plausibility for the effects of Pb on IgE is
provided by consistent findings in animals with gestational or gestational-lactational Pb exposures, with
some evidence at BLL relevant to humans. These findings are supported by strong evidence of Pb-
induced increases in T cell-derived helper (Th)2 cytokine production and inflammation in animals (U.S.
EPA. 2013).
Available toxicological evidence evaluated in the 2013 Pb ISA indicates a "likely to be causal"
relationship between Pb exposure and decreased host resistance. This conclusion was based primarily on
animal toxicological studies in which relevant Pb exposures decreased responses to antigens
(i.e., suppressed the delayed-type hypersensitivity (DTH) response and increased bacterial titers and
subsequent mortality in rodents). Further, evidence demonstrating biological plausibility, including
suppressed production of Thl cytokines and decreased macrophage function in animals support these
conclusions (U.S. EPA. 2013).
The 2013 Pb ISA also included an evaluation of the epidemiologic and toxicological evidence for
Pb-induced autoimmunity. Only a few toxicological studies provided evidence for Pb-induced generation
of autoantibodies and the formation of neoantigens that could result in the development of autoantibodies
following Pb exposure. Considering the limited evidence at hand, the available studies were inadequate to
determine if there is a causal relationship between Pb exposure and autoimmunity (U.S. EPA. 2013).
This ISA determined causality for adverse effects of Pb exposure on the three different aspects of
the immune system. Accounting for recent toxicological and epidemiologic studies demonstrating that Pb
exposure decreases host resistance to infection, suppresses the DTH response in animals, and decreases
the vaccine antibody response in children, there is sufficient evidence to conclude that there is likely to be
a causal relationship between Pb exposure and immunosuppression. Recognizing that recent
epidemiologic studies do not provide evidence of an association between exposure to Pb and atopic
disease and consistent toxicological evidence that exposure to Pb alters physiological responses in
animals consistent with allergic sensitization, the body of evidence supports changing the causal
determination from likely to be causal to suggestive of a causal relationship between Pb exposure and
sensitization and allergic responses. Evidence for effects of Pb exposure on autoimmunity and
autoimmune disease are disparate and highly limited. For that reason, the body of evidence describing the
relationship between exposure to Pb and autoimmunity remains inadequate to determine if a causal
relationship exists.
The following sections provide an overview of study inclusion criteria for this appendix
(Section 6.2), summaries of recent health effects evidence (Sections 6.3, 6.4, and 6.5), a discussion of
biological plausibility (Section 6.6), and a discussion of the causality determination for Pb exposure and
immune system effects (Section 6.7, Table 6-1, Table 6-2, and Table 6-3).
6-2
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6.2 Scope
The scope of this appendix is defined by Population, Exposure, Comparison, Outcome, and Study
Design (PECOS) statements. The PECOS statement defines the objectives of the review and establishes
study inclusion criteria, thereby facilitating identification of the most relevant literature to inform the Pb
ISA.1 In order to identify the most relevant literature, the body of evidence from the 2013 Pb ISA was
considered in the development of the PECOS statements for this appendix. Specifically, well-established
areas of research; gaps in the literature; and inherent uncertainties in specific populations, exposure
metrics, comparison groups, and study designs identified in the 2013 Pb ISA inform the scope of this
appendix. The 2013 Pb ISA used different inclusion criteria than the current ISA, and the studies
referenced therein often do not meet the current PECOS criteria (e.g., due to higher or unreported
biomarker levels). Studies that were included in the 2013 Pb ISA, including many that do not meet the
current PECOS criteria, are discussed in this appendix to establish the state of the evidence prior to this
assessment. With the exception of supporting evidence used to demonstrate the biological plausibility of
Pb-associated effects on the immune system, recent studies were only included if they satisfied all of the
components of the following discipline-specific PECOS statements:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect;
Exposure: Exposure to Pb2 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure3; or intervention groups in randomized trials and quasi-experimental studies;
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles);
Outcome: Immune system effects including but not limited to immunotoxicity, systemic
inflammation, and immune-based diseases; and
'The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
2Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area that was
relevant to the National Ambient Air Quality Standards review (e.g., longitudinal studies designed to examine recent
versus historical Pb exposure).
3Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 |im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 |im3 (PM2.5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is difficult
to assess the representativeness of these concentrations to population exposure [Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PM2 5 with BLLs are lacking.
6-3
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Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials, and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages);
Exposure: Oral, inhalation, or intravenous treatment(s) administered to a whole animal (in
vivo) that results in a BLL of 30 (ig/dL or below;4,5
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control;
Outcome: Immunological effects; and
Study Design: Controlled exposure studies of animals in vivo.
Consistent with this scoping, the following sections evaluate evidence for the effects of Pb
exposure on the immune system. In the 2013 Pb ISA, evidence for effects on the immune system was
organized into atopic and inflammatory responses, decreased host resistance, and autoimmunity.
Immunological evidence for this ISA is organized to reflect disease categories most relevant to Pb
exposure including immunosuppression (Section 6.3), sensitization and allergic responses (Section 6.4),
and autoimmunity and autoimmune diseases (Section 6.5). These categories encapsulate the immune-
related endpoints used in the 2013 Pb ISA while recognizing advances in the field of immunotoxicology.
The sections that follow focus on studies published since the completion of the 2013 Pb ISA. This
evidence is organized and weighed based on the World Health Organization's Guidance for
Immunotoxicity Risk Assessment for Chemicals (IPCS. 2012). As detailed in this guidance, data from
endpoints observed in the absence of an immune stimulus (e.g., levels of serum immunoglobulins, white
blood cell (WBC) counts, WBC differentials, T cell subpopulations, immune organ weights) are not
sufficient on their own to draw a conclusion regarding immune hazard but may provide useful supporting
evidence, especially when evaluated in the broader context of functional data (IPCS. 2012).
Consequently, the sections that follow are organized into two categories: the more informative measures
of immune system function and supporting immune system data. Study-specific details, including animal
4Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
5This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLLs.
The 95th percentile of the 2011-2016 National Health and Nutrition Examination Survey distribution of BLL in
children (1-5 years; n = 2,321) is 2.66 |.ig/dL (CDC. 2019) and the proportion of individuals with BLLs that exceed
this concentration varies depending on factors including (but not limited to) housing age, geographic region, and a
child's age, sex, and nutritional status.
6-4
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type, exposure concentrations, and exposure durations in experimental studies, and study design, exposure
metrics, and select results in epidemiologic studies are presented in evidence inventories in Section 6.8.
6.3 Immunosuppression
Immunosuppression can lead to the increased incidence and/or severity of infectious and
neoplastic diseases. Immunosuppressants may be identified using data generated from general toxicity
studies or through completion of dedicated immunotoxicity studies. In either case, evidence may be
collected from assays designed to assess the function of the immune system following xenobiotic
exposure or from supporting immune system endpoints.
6.3.1 Epidemiologic Studies of Immunosuppression
Epidemiologic studies relevant to immunosuppression generally include studies of viral and
bacterial infection and vaccine antibody response, as well as studies of WBCs and cytokines. A limited
number of epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA. 2013) provided evidence of
associations between cord blood or blood Pb and viral and bacterial infection in children. However, these
studies were cross-sectional and did not include adjustment for potential confounders, limiting the
strength of conclusions that could be drawn about the effects of Pb exposure on viral or bacterial
infections. Cross-sectional studies of cell-mediated immunity reported consistent associations between
BLL and lower T cell abundance in children, while results from other studies on lymphocyte activation,
macrophages, neutrophils, and natural killer (NK) cells were generally inconsistent or not sufficiently
informative (e.g., cross-sectional study designs with limited or no consideration of potential confounding
and a lack of information on concentration-response relationships).
There have been a number of recent epidemiologic studies of immunosuppression, including
prospective birth cohorts and studies with lower mean or median BLL than those reviewed in the 2013 Pb
ISA, many with measures of central tendency <2 (ig/dL. The recent studies also apply more robust
statistical methods and consistently consider a wider range of potential confounders. In general, recent
studies provide consistent evidence that exposure to Pb is associated with greater susceptibility to
infection and a less robust vaccine antibody response. Additionally, a group of studies in the same
population provides some evidence of altered immune cells and cytokines in association with BLLs.
Measures of central tendency for BLLs used in each study, along with other study-specific details,
including study population characteristics and select effect estimates, are highlighted in Table 6-4. An
overview of the recent evidence is provided below.
6-5
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6.3.1.1 Host Resistance
While the 2013 Pb ISA (U.S. EPA. 2013) evaluated a limited number of epidemiologic studies
that indicated an association between BLL and viral and bacterial infections in children, none of the
studies considered potential confounders and most analyzed populations with high mean BLLs (means
>10 (ig/dL). Recent studies expand the evidence base by examining populations with wider age-ranges
and much lower mean and median BLLs. The recent studies also adjust for a wide range of potential
confounders, including extensive consideration of SES factors.
Recent cross-sectional studies provide consistent evidence of associations between Pb exposure
and viral and bacterial infections, including Helicobacter Pylori, Toxoplasma Gondii, and hepatitis B
(Park et al.. 2020; Krueger and Wade. 2016). or susceptibility to antibiotic resistance measured via nasal
Staphylococcus aureus colonization (Eggers et al.. 2018). In a National Health and Nutrition Examination
Survey (NHANES) analysis including children and adults, each 1 (ig/dL higher level of blood Pb was
associated with 8 to 10% higher odds of H. Pylori (odds ratio [OR]: 1.09 [95% confidence interval (CI):
1.05, 1.13]), T. Gondii (OR: 1.10 [95% CI: 1.06, 1.14]), and hepatitis B (OR: 1.08 [95% CI: 1.03, 1.13])
seropositivity in the U.S. population (Krueger and Wade. 2016). Positive associations were persistent, but
varied in magnitude across more specific age groups, including children under 13, participants aged 13 to
35, and adults >35 years old. The associations for H. Pylori were markedly stronger in magnitude for
children less than 13 years old compared with the other age groups, whereas the associations for T.
Gondii were slightly weaker in children. Additionally, in multipollutant models with cadmium (Cd), there
was no evidence to suggest additive or multiplicative interaction between Pb and Cd. Another cross-
sectional study of adults with abnormal lesions identified during endoscopy also reported that H. Pylori
infection rates were associated with increased BLL (Park et al.. 2020).
In addition to cross-sectional studies, a recent test-negative case-control study reported that peak
BLLs were associated with elevated influenza and respiratory syncytial virus (RSV) rates in children
<4 years old presenting with relevant symptomology (Feiler et al.. 2020). Test-negative case-control
study designs are often used in vaccine efficacy studies to control for healthcare seeking behaviors, but
for the intended purposes of this study, the design could bias results toward the null if the non-RSV and
influenza illnesses are also related to Pb-induced immune deficiencies. The results in the full population
were adjusted for fewer potential confounders (i.e., age, sex, race, ethnicity, insurance status, and season)
on account of missing variables, and the observed associations were null in a notably reduced sample
population (<25%) with expanded adjustment for confounders.
6.3.1.2 Antibody Responses
There were no studies evaluated in the 2013 Pb ISA (U.S. EPA. 2013) that examined the
relationship between exposure to Pb and vaccine antibody response in children. There are a few recent
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studies that provide generally consistent evidence of Pb-related decreases in vaccine antibodies in
populations with low mean or median BLL.
In a birth cohort of vaccinated children in South Africa, Di Lenardo et al. (2020) reported that
each 1 (ig/dL higher level of blood Pb at age 1 was associated with 13% (95% CI: 2%, 26%) higher risk
of tetanus IgG titers below the protective limit at age 3.5 years. A key strength of this study was its
prospective nature and the timing of blood Pb measures that approximately coincided with vaccine
administration. The authors also examined measles and Haemophilus influenzae type B (Hib) IgG levels
but did not observe associations with BLL. Cross-sectional studies—including a large NHANES analysis
of children ages 6 to 17 years old (Jusko et al.. 2019) and another small study comparing kindergarten-
aged children in China living near an e-waste facility to those in a nearby community with similar
sociodemographic characteristics (Xu et al.. 2015)—also provide evidence that higher BLLs are
associated with lower counts of virus-neutralizing antibodies. However, unlike the results from Di
Lenardo et al. (2020). Jusko et al. (2019) reported that higher BLLs were associated with lower counts of
anti-measles IgG antibodies, as well as anti-mumps antibodies. The authors observed a null association
with anti-rubella IgG levels. In the analysis in China, Xu et al. (2015) noted that geometric mean BLL
dropped precipitously between the 2 years of the study (>3 (.ig/dL). The authors conducted an analysis
stratified by the year of the study and observed lower anti-hepatitis B surface antigen (HBsAb) titers in
relation to higher BLLs in both years; however, the association was notably stronger in magnitude in the
year with higher geometric mean BLL (2011: -0.447 s/co [95% CI: -0.491, -0.403 s/co]; 2012: -0.366
s/co [95% CI: -0.404, -0.328 s/co] per 1 (ig/dL higher BLL). An important uncertainty in this analysis is
potential confounding by other contaminants present in the community. In contrast to the previously
discussed evidence, a birth cohort of vaccinated children in Bangladesh reported a positive association
between cord BLL and diphtheria and tetanus IgG antibodies at age 5 (Welch et al.. 2020). Notably, the
associations were null when the exposure metric was concurrent BLL rather than cord blood Pb.
6.3.1.3 White Blood Cells and Cytokines
Several epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA. 2013) examined the
relationship between Pb exposure and changes in WBC populations (i.e., counts and phenotypes) and
cytokine levels. Although WBC counts and cytokine levels are commonly evaluated in epidemiologic
studies, these data can be challenging to interpret because (1) WBC populations are not particularly
sensitive indicators of immunotoxicity and (2) changes in cytokine levels can be associated with many
different types of tissues and toxicities, either as part of cell differentiation to different immune cell types
or including site-specific inflammation, which reflects an immune response to tissue injury but not
necessarily an effect on or impairment of immune function (Tarrant. 2010). For these reasons, WBC
populations and cytokine secretion data (in the absence of a stimulus) are not considered apical outcomes
for the purpose of identifying immune hazard, but rather as supporting evidence for understanding
mechanisms of immune disruption.
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In the 2013 Pb ISA, there was generally consistent evidence of positive associations between
BLLs and T cell counts in children, but epidemiologic evidence for other immune cell and cytokine
measures were uninformative due to cross-sectional study designs with limited or no consideration of
potential confounding and a lack of information on the concentration-response relationship. Recent
studies provide some evidence of associations between Pb exposure and immune cell and cytokine
abundance in children, though the number of studies examining overlapping immunological markers is
limited.
The majority of recent epidemiologic studies of WBCs and cytokines come from a group of
related, small cross-sectional studies evaluating a study population of kindergarten-aged children in
Guangdong, China living either near an e-waste facility or in a nearby community with otherwise similar
sociodemographic characteristics and pollutant exposures (Chen et al.. 2021; Zhang et al.. 2020; Huo et
al.. 2019; Cao et al.. 2018; Dai et al.. 2017). Across these studies, authors reported that BLLs were
positively associated with a number of biomarkers related to immunological function, including the
proinflammatory cytokines interleukin (IL)-1|3 (Zhang et al.. 2020; Huo et al.. 2019). IL-12p70, and
interferon (IFN)-y (Huo et al.. 2019) and pleiotropic cytokine IL-6 (Zhang et al.. 2020). Chronic
inflammation has the potential to contribute to the development of immunosuppression (Kanterman et al..
2012). In addition, higher BLLs were associated with differences in several other biomarkers of immune
system function including higher erythrocyte complement receptor type 1 (CR1) expression (Dai et al..
2017); a higher percentage of cluster of differentiation (CD)4+ central memory T cells (Cao et al.. 2018);
higher neutrophil counts (Zhang et al.. 2020); higher counts of WBCs, neutrophils, and monocytes (Chen
et al.. 2021); a lower percentage of CD4+ naive T cells (Cao et al.. 2018); and lower levels of tumor
necrosis factor alpha (TNF)-a (Zhang et al.. 2020). The authors of these studies also noted some null
associations with BLLs, including CD3+, CD4+, and CD8+ cell counts (Cao et al.. 2018) and monocytes,
lymphocytes, IL-8, and IL-10 (Zhang et al.. 2020). Consistent with Chen et al. (2021). another cross-
sectional study in China with a similar design (e.g., kindergartners recruited from reference and control
communities with and without industrial exposure to Pb) reported null associations between BLL and
odds of lowerlower WBC counts (Li et al.. 2018).
In the only recent study of an adult population, a small cross-sectional analysis of oil spill
response workers with low BLLs (mean: 1.82 (.ig/dL). Werder et al. (2020) observed associations between
higher BLLs and higher proinflammatory cytokines (i.e., IL-1|3 and IL-8) and pleiotropic cytokine IL-6
but not the proinflammatory cytokine TNF-a. This was generally consistent with the previously discussed
results in children, with the exception of IL-8 for which a null association was reported in children.
Notably, as highlighted in a stratified analysis to examine effect modification by obesity, the observed
associations are entirely driven by associations in obese participants Werder et al. (2020). For example,
each 1 (ig/dL higher level of blood Pb was associated with 72.8 pg/mL (95% CI: 36.9, 108.7 pg/mL)
higher IL-6 levels in the entire study population. However, in the stratified analysis, the association was
stronger in magnitude in obese participants (169.6 pg/mL [95% CI: 119.8, 219.4 pg/mL]) and null in non-
obese participants (-2.6 pg/mL [95% CI: -45.5, 40.3 pg/mL]).
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6.3.2
Toxicological Studies of Immunosuppression
Toxicological studies evaluated in the 2013 Pb ISA (U.S. EPA, 2013) investigating Pb-induced
immunosuppression were derived from several lines of evidence including functional assays (i.e., host
resistance, antibody responses, DTH response, and ex vivo WBC function) and bolstered by various
forms of supporting immune system data. Some of these data were reviewed in the 2006 Pb Air Quality
Criteria Document (AQCD) (U.S. EPA, 2006). Based on these previous evaluations, there is clear
evidence that exposure to Pb decreases host resistance to bacterial infection and increases production of
some pathogen-specific antibody subtypes promoting the shift toward Th2-type immune responses. The
results of investigations of the T cell dependent antibody response were inconsistent, with one study
reporting a decrease in the antibody response (BLL not reported) and another showing no effect in mice
with high BLLs (i.e., 59-132 (.ig/dL). However, Pb has consistently been shown to suppress the DTH
response in animal models. The DTH assay has a long history of use in immunotoxicity testing and i s
considered one of the most predictive immunotoxicity tests available (Dietert et al., 2010). Suppression
of the DTH response is a hallmark of Pb exposure. Exposure to Pb suppressed the DTH response in rats
(Chen et al., 2004; Bunn et al„ 2001a; Bunn et al„ 2001b; Chen et al„ 1999; Miller etal., 1998) and
chickens (Lee et al., 2002; Lee et al., 2001) with BLLs relevant to this ISA (i.e., <30 (ig/dL). The DTH
response was also suppressed by Pb exposures in other studies not reporting a BLL (Laschi-Loquerie et
al., 1984; Faith et al., 1979; Miilleretal., 1977) and in studies reporting BLLs outside the scope of this
ISA (Bunn et al„ 2001c; McCabe et al., 1999; Fandrich et al„ 1995).
Pb exposure also affected the functions of various WBCs under ex vivo conditions leading to (1)
suppression of Thl-mediated immunity (i.e., suppressed Thl cytokine production (e.g., IFN-y) and DTH
response); (2) altered macrophage function (e.g., increased reactive oxygen species [ROS] production,
decreased nitric oxygen [NO] production); and (3) reduced monocyte/macrophage phagocytosis. In
addition to assessing the effect of Pb on measures of immune system function, the effects of Pb exposure
on various immunotoxicology-related supporting immune system endpoints were also evaluated,
including (1) total serum immunoglobulins, (2) immune organ weight, (3) WBC number in the spleen,
thymus, lymph nodes, and bone marrow, and (4) WBC counts and subpopulation data collected from
blood samples. Generally, the number of these studies was limited and differences in study design and the
specific endpoints measured create challenges when interpreting these supporting immune system data.
Recent toxicological studies are limited in number and report on disparate outcomes, but
generally support evidence reported in the 2013 Pb ISA. Consistent with findings reported in the 2013 Pb
ISA, Pb exposure was again shown to suppress the DTH response. There are no recent toxicology studies
investigating the effects of Pb exposure on host resistance; however, there is some recent evidence that Pb
exposure altered the levels of some classes of antigen-specific antibodies in iron-deficient rats. Pb
exposure also reduced the total serum levels of some immunoglobulins in rats. As with the 2013 Pb ISA,
the effects of Pb on immune organ pathology and spleen weight were inconsistent. New to this ISA, a
recent study reported that Pb exposure decreased relative thymus weight. Differences in experimental
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design and the specific types of WBCs assessed complicate interpretation of data collected on the number
and relative abundance of the different types of WBCs in the spleen, thymus, lymph nodes, and bone
marrow following exposure to Pb. WBC counts and subpopulation data collected from hematological
investigations are similarly challenging to interpret.
6.3.2.1 Host Resistance
Available toxicological evidence evaluated in the 2013 review provides clear evidence that host
resistance to bacterial infection is compromised following Pb exposures, resulting in BLLs as low as
20 (ig/dL. The 2013 Pb ISA (U.S. EPA. 2013) reported several rodent host resistance studies wherein
mortality was increased in pathogen-exposed animals that were also exposed to Pb through drinking
water. For example, various studies reported decreased clearance of bacteria and increased mortality
induced by Listeria monocytogenes in mice exposed postnatally to Pb acetate in drinking water for 3 to
8 weeks, resulting in BLLs ranging from 20-25 (ig/dL (Dvatlov and Lawrence. 2002; Kim and Lawrence.
2000; Kishikawa et al.. 1997; Lawrence. 1981). Other studies reported increased mortality from
Salmonella or Escherichia coli, or decreased clearance of Staphylococcus, in mice administered Pb
acetate or Pb nitrate via injection, resulting in BLLs relevant to the 2013 Pb ISA (Fernandez-Cabezudo et
al.. 2007; Bishavi and Sengupta. 2006; Cook et al.. 1975; Hemphill et al.. 1971; Selve et al.. 1966). In
addition to high BLL (i.e., 71-313 (.ig/dL). increased mortality from viral infection was also reported in
mice and chickens administered Pb (mostly Pb acetate) for 4-10 weeks (Gupta et al.. 2002; Exon et al..
1979; Thind and Khan. 1978). Further, evidence suggested a plausible mode of action involving
suppressed production of Thl cytokines (Fernandez-Cabezudo et al.. 2007; Lara-Tcicro and Panicr.
2004). decreased macrophage function (Lodi et al.. 2011; Bishavi and Sengupta. 2006; Chen et al.. 1997;
Hilbertz et al.. 1986; Castranova et al.. 1980). and increased inflammation in animals (Miller et al.. 1998;
Bavkov et al.. 1996; Zelikoff et al.. 1993).
There were no recent toxicology studies investigating the effects of Pb exposure on host
resistance that satisfied the PECOS criteria described in Section 6.2 available for this review.
6.3.2.2 Delayed-Type Hypersensitivity Responses
Antigen-specific cell-mediated immune (CMI) responses are a key component of host defense
mechanisms against virally infected cells, tumor cells, and certain fungal infections. The DTH assay is a
standard test for assessing CMI responses in animals (IPCS. 2012). As noted in the 2013 Pb ISA,
suppressed DTH response is one of the most consistently reported immune effects associated with Pb
exposure in animals (U.S. EPA. 2013). Suppression of the DTH response has been reported following
gestational (Chen et al.. 2004; Bunn et al.. 2001a; Bunn et al.. 2001b. c; Lee et al.. 2001; Chen et al..
1999; Miller etal.. 1998; Faith etal.. 1979) and postnatal (McCabe et al.. 1999; Laschi-Loquerie et al..
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1984; Miilleretal.. 1977) exposures to Pb acetate resulting in BLLs ranging from 6.75 to >100 (ig/dL) in
rats, mice and chickens (U.S. EPA. 2013).
In a recent study, administration of Pb acetate in drinking water for 42 days (BLL = 18.48 (ig/dL)
significantly suppressed the DTH response in adult male Sprague Dawley rats (Fang et al.. 2012). To
explore the role of regulatory T cells (Tregs) in the DTH response, Fang et al. (2012) employed a T cell
transfer model. Total CD4+ T cells and CD4+CD25- cells were collected from control and Pb-exposed
rats and then transferred to recipient rats that were subsequently challenged with l-Fluoro-2,4-
dinitrobenzene (DNFB) to induce a DTH response. The DTH response was diminished in rats receiving
CD4+ T cells from Pb-exposed rats compared with those receiving CD4+ cells from control animals.
Importantly, the effect was lost when Tregs were depleted from the pool of CD4+ cells transferred to the
recipient rats. These findings suggest that Tregs play a critical role in Pb-induced immune suppression
(Fang et al.. 2012). Study-specific details, including animal species, strain, sex, and BLLs are highlighted
in Table 6-6.
6.3.2.3 Antibody Responses
The production of antigen-specific antibodies is a major defense mechanism of humoral immune
responses. Only one study reporting effects on antigen-specific antibody responses was evaluated in the
2013 Pb ISA (U.S. EPA. 2013). In that study, Fernandez-Cabezudo et al. (2007) reported no difference in
the serum levels of Salmonella-specific IgM following infection with a sublethal dose of Salmonella
(1.5 x 104 organisms/mouse) in control C3H/HeN mice and mice exposed to 10 mM Pb acetate in
drinking water for 16 weeks (resultant mean BLL: 106 (.ig/dL). However, compared with control mice,
mice exposed to Pb acetate had less IgG2a and more IgGl antibodies providing evidence for a shift
toward Th2-type immune responses resulting in decreased resistance to Salmonella enterica (Fernandez-
Cabezudo et al.. 2007). Studies describing effects of Pb exposure on the T cell dependent antibody
response (TDAR) were also reviewed in the 2013 Pb ISA. The TDAR is a comprehensive immune
function assay that integrates several aspects of immune responses. Thus, xenobiotic-induced alterations
in antigen processing and presentation, B and T cell interactions, antibody production, and isotype class
switching and modification have the potential to modify this defense mechanism (IPCS. 2012). Results of
the TDAR response to sheep RBCs have been inconsistent. For example, the TDAR was significantly
decreased in mice exposed to Pb acetate through drinking water for 3 weeks, resulting in BLLs of
25.4 (ig/dL (Blaklev and Archer. 1981). However, in a second drinking water study, the TDAR was
increased in 1 of 8 mouse strains (the other 7 strains were unaffected) evaluated following administration
of Pb acetate in drinking water for 8 weeks resulting in high BLL (mean range 59-132 (ig/dL) (Mudzinski
et al.. 1986).
In a recent study, adult Sprague Dawley rats (data from both sexes pooled) were fed either a
control diet or an iron-deficient diet for the duration of the experiment (Yathapu et al.. 2020). After
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confirming iron deficiency at 4 weeks, rats were administered Pb acetate in drinking water for 4 weeks.
At this time, a subset of mice was vaccinated with tetanus toxoid (TT). Rats received two booster doses
(2-week interval) before assessing antigen-specific antibody levels 2 weeks after the last booster dose.
Under these conditions, Pb acetate (BLL = 16.1 (ig/dL) had no effect on the levels of anti-TT-specific IgG
and IgM antibodies in the serum of rats that received the control diet whereas the levels of anti-TT-
specific IgM were decreased and those of IgG were unaffected in the serum of iron-deficient rats
(Yathapu et al.. 2020). Study-specific details, including animal species, strain, sex, and BLLs are
highlighted in Table 6-5.
6.3.2.4 Ex Vivo White Blood Cell Function
White blood cells are cells of the immune system involved in protecting the body from infectious
disease. These cells can be organized into two lineages—myeloid cells and lymphoid cells. Myeloid cells
(i.e., myelocytes) include neutrophils, eosinophils, mast cells, basophils, and monocytes. Lymphoid cells
(i.e., lymphocytes) include T cells, B cells, and NK cells. Xenobiotic-induced alterations in ex vivo WBC
function is considered clear evidence of immunosuppression (IPCS, 2012). Ex vivo WBC function assays
are performed outside the body using immune cells collected from exposed individuals.
The 2013 Pb ISA reviewed the effects of Pb exposure on the functions of various WBCs under ex
vivo conditions indicating (1) a shift in lymphocyte cytokine production towards the production of Th2
cytokines (Heo et al., 2007; McCabe and Lawrence, 1991), reduced number of Thl cells and Thl
cytokine levels (McCabe and Lawrence, 1991), (2) increased dendritic cell induced Th2 cell proliferation
and cytokine production (Gao et al., 2007), and (3) reduced monocyte/macrophage phagocytosis (Lodi et
al., 2011; Bussolaro et al„ 2008; Deng and Poretz, 2001; Kowolenko et al., 1991; Zhou et al., 1985) and
decreased NO production (Farrer et al., 2008; Mishraet al., 2006; Bunn et al., 2001b; Lee et al., 2001;
Krocova et al., 2000; Chen et al., 1997; Tian and Lawrence, 1996; Tian and Lawrence, 1995). No studies
on neutrophils and NK cells were reviewed in the 2013 Pb ISA.
A few PECOS-relevant papers evaluating the effects of Pb exposure on ex vivo WBC function
have been published since the 2013 Pb ISA. Fang et al. (2012) reported that administration of Pb acetate
in drinking water for 42 days (BLL = 18.48 (ig/dL) had no effect on the suppressive properties of Tregs
isolated from adult male Sprague Dawley rats. In a second study, the effects of Pb administration on
Concanavalin A (Con A)-stimulated lymphocyte proliferation and cytokine production were investigated
(Yathapu et al„ 2020). For this investigation, adult male and female Sprague Dawley rats were fed either
a control diet or an iron-deficient diet for the duration of the experiment. After confirming iron deficiency
at 4 weeks, the rats were administered Pb acetate in drinking water for 4 weeks. At this time, a subset of
rats was vaccinated with TT. Rats received two booster doses (2-week interval) before splenocytes were
collected 2 weeks after the last booster dose. Irrespective of vaccine status, Pb treatment
(BLL =16.1 (ig/dL) had no effect on Con A-stimulated proliferation of splenocytes collected from rats
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fed the control diet. However, when rats were fed an iron-deficient diet, Pb treatment (BLL = 41.6 (ig/dL)
increased Con A-stimulated splenocyte proliferation (Yathapu et al.. 2020). Unfortunately, because of
incomplete reporting, data related to cytokine production by Con A-stimulated splenocytes reported by
Yathapu et al. (2020) are not interpretable. In addition, Cai et al. (2018) measured cytokine levels directly
in blood and reported that, administration of Pb acetate drinking water (0.2%; BLL = 9.3 (ig/dL) for
84 days had no effect on erythropoietin, granulocyte-macrophage colony-stimulating factor (GM-CSF),
interleukin (IL)-6, and TNF-a levels in adult Sprague Dawley rats (data from sexes pooled). Study-
specific details, including animal species, strain, sex, and BLLs are highlighted in Table 6-7 and
Table 6-14.
6.3.2.5 Immune Organ Pathology
The 2013 Pb ISA did not report on the effects of Pb exposure on immune organ pathology (U.S.
EPA. 2013). However, xenobiotic exposure can alter primary immune sites important for immune cell
maturation, including the bone marrow, liver, thymus, and Peycr's patches. Secondary lymphoid sites
(i.e., spleen, lymph nodes, tonsils) can also be affected by exposure to immunotoxicants. Data from these
endpoints are not sufficient on their own to draw a conclusion regarding immune hazard, but may provide
useful supporting evidence (IPCS. 2012). Pb-induced alterations in immune organ pathology were not
addressed in the 2013 Pb ISA.
Since the 2013 Pb ISA, there have been three reports published that included an assessment of
immune organ pathology following exposure to Pb and that fit the PECOS criteria described in
Section 6.2. In the first study, Pb treatment induced changes in the spleen architecture of adult male
C57BJ mice exposed via drinking water (200 ppm; BLL = 21.6 (ig/dL) for 45 days. These changes
included increasing the amount of white pulp (qualitative) and decreasing the definition of the
germinative center of the inner peri-arteriolar lymphoid sheath, but the marginal zone was unaffected
(Corsetti et al.. 2017). In a different study, inhalation of Pb oxide nanoparticles (1.23 x 106 x 10
particles/cm3, 24 hours/day for 6 weeks BLL 13.9 (ig/dL) had no effect on spleen pathology in two
experiments conducted in adult female Institute for Cancer Research (ICR) mice (Dumkova et al.. 2017).
Dumkova et al. (2020a) conducted another study with Pb oxide nanoparticles (68.6 x 106 particles/cm3,
24 hours/day for up to 6 weeks) in CD-l(ICR) mice that included histological analysis of the spleen, but
did not report their findings. Exposure to Pb oxide nanoparticles (0.95 6 x 106 particles/cm3, 24 hours/day
for 11 weeks, BLL =18.1 (ig/dL) had no effect on spleen histopathology in CD-I (ICR) BR mice (Smutna
et al.. 2022). Study-specific details, including animal species, strain, sex, and BLLs are highlighted in
Table 6-8.
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6.3.2.6
Immunoglobulin Levels
Immunoglobulins (i.e., antibodies) are produced by plasma cells (i.e., differentiated B cells).
Immunoglobulins are a critical part of the immune response and act by recognizing and binding to
specific antigens such as bacteria and viruses leading to their destruction. Although immunoglobulin type
and quantity are easy to measure in serum, their levels are difficult to interpret in the absence of a
controlled immune challenge. For this reason, these data are not considered a predictive measure for
immunotoxicity and are most useful for supporting data collected from immune functional assays. The
2013 Pb ISA reviewed the effects of Pb exposure on total serum IgE in the context of immediate-type
hypersensitivity (Chen et al., 2004; Snyder et al., 2000; Miller etal., 1998; Heo et al., 1997; Heo et al.,
1996). In addition, the 2013 Pb ISA reviewed the effects of Pb exposure on total serum IgG subtypes
(Kasten-Jolly et al., 2010; Carey et al., 2006; Gao et al., 2006; Snyder et al., 2000). While noting that the
BLLs were not relevant to human exposures, the 2013 Pb ISA described the observed effects as
inconsistent.
Since the 2013 Pb ISA, only one PECOS-re levant publication included an assessment of total
serum immunoglobulin levels following exposure to Pb. For this investigation, adult Sprague Dawley
(data from sexes pooled) were fed either a control diet or an iron-deficient diet for the duration of the
experiment. After confirming iron deficiency after 4 weeks, rats were administered Pb acetate in drinking
water for 4 weeks. At this time, a subset of mice was vaccinated with TT. Rats received two booster doses
(2-week interval) before splenocytes were collected 2 weeks after the last booster dose. Irrespective of
vaccine status, Pb treatment reduced mucosal IgA levels in rats fed the control diet (BLL =16.1 (ig/dL).
Under conditions of iron deficiency, Pb treatment further reduced mucosal IgA levels
(BLL = 41.6 (ig/dL). Total serum IgM and IgG were unchanged by Pb under all conditions evaluated
(Yathapu et al„ 2020). Study-specific details, including animal species, strain, sex, and BLLs are
highlighted in Table 6-9.
6.3.2.7 Immune Organ Weights
Changes in lymphoid organ weights (thymus, spleen, lymph node, or bone marrow) may indicate
immunotoxicity and are useful for supporting data collected on immune function. As reported in the 2013
Pb ISA, exposure to Pb increased relative spleen weight in mice and rats exposed to Pb acetate and Pb ion
in drinking water (U.S. EPA. 2013). In the only available study, lymph node weight decreased following
exposure to Pb acetate (Institoris et al.. 2006). There were no studies that evaluated changes in thymus
weight reviewed in the 2013 Pb ISA. Several recent studies evaluating the effects of Pb exposure on
lymphoid tissues are described below, including one study describing effects on the thymus. Study-
specific details, including animal species, strain, sex, and BLLs are highlighted in Table 6-10.
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6.3.2.7.1
Thymus Weight
The thymus, which is essential for T cell development, is a critically important component of the
immune system; changes in thymus weight are a more sensitive indicator of immunotoxicity than changes
in spleen weight. Relative thymus weight was significantly decreased in juvenile Sprague Dawley rats
(data from sexes pooled) orally administered Pb acetate (1 or 10 mg/kg with BLL of 3.27 (ig/dL and
12.5 (ig/dL, respectively) for up to 25 days (Graham et al.. 2011). A second study performed by the same
laboratory using the same experimental design investigated the effects of oral administration of Pb acetate
(gavage) on relative thymus weight (Amos-Kroohs et al.. 2016). Because of incomplete reporting,
however, the effect of Pb on thymus weight could not be discerned and this element of the study was
rejected for study quality deficiencies.
6.3.2.7.2 Spleen Weight
The spleen has a prominent role in immune function, as well as serving as a reservoir for
monocytes. The effect of Pb administration via oral and inhalation routes in rats and mice has been
recently investigated. In juvenile Sprague Dawley rats (data from sexes pooled), relative spleen weight
was not affected following oral administration of Pb acetate (gavage, 1 or 10 mg/kg with BLL up to 3.27
and 12.5 (ig/dL, respectively) for up to 25 days (Amos-Kroohs et al.. 2016; Graham et al.. 2011).
Absolute spleen weight, however, was decreased significantly following exposure to 10 mg/kg
(BLL = 12.5 (ig/dL) Pb acetate (Graham et al.. 2011). Similarly, spleen weight was unaffected in adult
male Wistar rats exposed to Pb acetate in drinking water (357 (ig/kg/day or 1607 (ig/kg/day with BLL of
1.77 ± 0.7 (ig/dL and 8.6 ± 2.9 (ig/dL, respectively) for 4 weeks (Wildemann et al.. 2015). In the only
study investigating the effects of Pb exposure in mice, Pb acetate treatment significantly increased
relative spleen weight in adult male C57BJ mice exposed via drinking water (200 ppm,
BLL = 21.6 (ig/dL) for 45 days (Corsetti et al.. 2017).
Effects of Pb exposure through inhalation were inconsistent. Inhalation exposure to Pb oxide
nanoparticles (1.23 x 106 nanoparticles/cm3, BLL 13.9 (ig/dL) increased relative spleen weight in adult
female ICR mice exposed for 6 weeks, but the finding was not replicated in a duplicate experiment
performed as part of the same study (Dumkova et al.. 2017). In a second study performed by the same
lead investigator, inhalation exposure to a higher concentration of Pb oxide nanoparticles (2.23 x 106
nanoparticles/cm3) for a longer duration (i.e., 11 weeks) had no effect on relative spleen weight adult
female CD-l(ICR) BR mice with a BLL of 17.4 (ig/dL (Dumkova et al.. 2020b). However, inhalation
exposure to Pb (II) nitrate nanoparticles (68.6 x 106 nanoparticles/cm3) decreased relative spleen weight
in adult female CD-l(ICR) BR mice exposed for 2 weeks (BLL = 4.0 (.ig/dL). but the effect was not
observed at the 6 week or 11-week timepoints with BLL up to 8.5 (ig/dL (Dumkova et al.. 2020a).
Similarly, exposure to Pb oxide nanoparticles (0.956 x 106 particles/cm3, 24 hours/day for 11 weeks,
BLL =18.1 (ig/dL) had no effect on relative spleen weight in CD-I (ICR) BR mice (Smutna et al.. 2022)
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6.3.2.8 White Blood Cell Counts and Differentials (Spleen, Thymus, Lymph node,
Bone Marrow)
Changes in WBC number and differentials collected from lymphoid organs may indicate
immunotoxicity and are useful for supporting data collected from immune function assays. Although
there were no data for WBC counts and differentials in lymphoid tissues reviewed in the 2013 Pb ISA,
several recent studies describing the effects of Pb exposure on lymphoid tissues are described below.
Study-specific details, including animal species, strain, sex, and BLLs are highlighted in Table 6-11.
6.3.2.8.1 Spleen
The effects of Pb exposure on spleen cellularity were investigated in three recent studies.
Administration of Pb acetate in drinking water (300 ppm; BLL = 18.48 (ig/dL) for 42 days significantly
increased the number of Tregs, reduced the absolute number of CD3+ cells and the percentage of CD4+ T
cells, but not the percentage CD8+ T cells in the spleens of adult male Sprague Dawley rats (Fang ct al..
2012). In contrast, administration of Pb acetate in drinking water for 28 days had no effect on percentage
of CD4+ cells, but the percentage of CD8+ cells was significantly increased in the spleens of adult male
and female Sprague Dawley rats (BLL =16.1 (ig/dL) (Yathapu et al.. 2020). Drinking water exposure to
Pb acetate (1250 ppm; BLL 4.7-41.3 (ig/dL) for 56 days decreased the number of innate lymphoid cells
(ILC), type 1 innate lymphoid cells (ILC1), NK- like ILC1 (NK-ILC1), type 2 innate lymphoid cells
(ILC2), and type 3 innate lymphoid cells (ILC3), but Pb had no effect on cell proliferation in vivo in
spleens collected from adult male and female (samples pooled) C57BL/6 mice (Zhu et al.. 2020).
6.3.2.8.2 Thymus
Pb acetate treatment had no effect on the total number of thymocytes or the number of thymic
CD4-/CD8- and CD4+CD8+ cells, but reduced the number of thymic CD4+CD8- cells by 25% and
slightly increased the number of CD4-CD8+ cells in adult male Sprague Dawley rats exposed via
drinking water (300 ppm; BLL = 18.48 (ig/dL) for 42 days (Fang et al.. 2012). Administration of Pb in
drinking water (300 ppm) for 42 days resulted in a 1.59-fold increase in the number of Tregs in the
thymus of adult male Sprague Dawley rats exposed (Fang et al.. 2012). There are no other recent studies
meeting PECOS criteria available for this endpoint.
6.3.2.8.3 Lymph Node
Two recent studies investigated the effects of Pb exposure on lymph node cellularity.
Administration of Pb acetate in drinking water (300 ppm; BLL = 18.48 (ig/dL)) to adult male Sprague
Dawley rats for 42 days had no effect on the absolute number of CD8+ T cells but reduced the absolute
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number of CD3+ cells and CD4+ T cells and increased the number of Tregs in the lymph nodes (type not
specified) (Fang et al.. 2012). Drinking water exposure to Pb acetate (1250 ppm; BLL 4.7-41.3 (ig/dL)
for 56 days decreased the number of ILCs, ILCls, NK-like ILCls (NK-ILCls), ILC2s, and ILC3s in
cervical lymph nodes collected from adult male and female (samples pooled) C57BL/6 mice (Zhu et al..
2020).
6.3.2.8.4 Bone Marrow
Two recent studies investigated the effects of Pb exposure on populations of immune cells in
bone marrow. Administration of Pb acetate in drinking water (0.2%; BLL = 9.3 (ig/dL) for 84 days had no
effect on the number of CD90+CD45- pluripotent hematopoietic stem cells in bone marrow collected
from adult male and female Sprague Dawley rats (Cai et al.. 2018). In a second study, administration of
Pb acetate in drinking water (1250 ppm; BLL 4.7-41.3 (ig/dL) for 56 days decreased the number of ILC
progenitors (ILCPs) and reduced number of ILCPs in the bloods of adult C57BL/6 mice (data from sexes
pooled) (Zhu et al.. 2020). These data suggest that Pb exposure impaired mobilization of ILCP cells to the
periphery. In the same study, the number of ILCs, ILCls, NK-ILCls, ILC2s, and ILC3s in bone marrow
were reduced, but Pb had no effect on cell proliferation in vivo (Zhu et al.. 2020). Pb suppressed
proliferation of ILCP in bone marrow, however.
To determine if the increase in the number of ILCPs associated with Pb exposure was caused by
impeded differentiation, common lymphoid progenitors from the bone marrow of Pb-treated (1250 ppm,
56 days; BLL 4.7-41.3 (ig/dL) or control enhanced green fluorescent protein (EGFP) mice were
transplanted into Pb-treated or control B6 mice (Zhu et al.. 2020). Common lymphoid progenitors
collected from Pb-treated EGFP mice gave rise to more ILCs compared with common lymphoid
progenitors from control donors in both Pb-treated and control recipients. Furthermore, common
lymphoid progenitors from Pb-treated donors produced more mature ILCs in control recipients than in
Pb-treated recipients. These findings indicate that common lymphoid progenitors in Pb-treated mice could
differentiate into mature ILCs, however, the Pb-treated host environment impeded differentiation into
ILCPs.
6.3.2.9 White Blood Cell Counts (Hematology and Subpopulations)
Changes in WBC number and differentials in blood may indicate potential immunotoxicity and
are useful for supporting data collected on immune function. The 2013 Pb ISA reviewed one toxicology
study that described the effects of Pb exposure on WBC numbers in blood (Sharma et al.. 2010). In that
study, the total number of WBCs, lymphocytes and monocytes were reduced in male Swiss albino mice
treated with Pb nitrate (50 mg/kg/day) (Sharma et al.. 2010). The effect of Pb exposure on WBC counts
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and subpopulations in blood reported in four recent studies are described below. Study-specific details,
including animal species, strain, sex, and BLLs are highlighted in Table 6-12.
Administration of Pb acetate in drinking water (0.2%, BLL 30.9 ± 14.7 (ig/dL) for 1 day had no
effect on the number of WBC, lymphocytes and neutrophils in whole blood collected from adult male
Wistar rats (Andielkovic et al.. 2019). However, when Pb acetate was administered in drinking water
(200 ppm; BLL = 21.6 (ig/dL) for 45 days consecutively, the numbers ofWBCs, neutrophils,
lymphocytes, and eosinophils decreased while the numbers of monocytes and basophils were unchanged
in blood collected from adult male C57BJ mice (Corsetti et al.. 2017). Changes in WBC number and
subpopulations were reported in a second study wherein the total number ofWBCs and the number of
CD4+ and CD8+ T cells were reduced in blood collected from male and female Sprague Dawley rats
(data from sexes pooled) following exposure to Pb acetate in drinking water (0.2%; BLL = 9.3 (ig/dL) for
84 days (Cai et al.. 2018). Additionally, exposure to Pb acetate (drinking water, 1250 ppm, BLL 4.7-
41.3 (ig/dL) for 56 days decreased the number of ILCs, type 1 innate lymphoid cells (ILC1), NK-like
ILC1 (NK-ILC1), type 2 innate lymphoid cells (ILC2), and type 3 innate lymphoid cells (ILC3). Pb
exposure additionally suppressed proliferation of ILCP in blood collected from adult male and female
(samples pooled) C57BL/6 mice (Zhu et al.. 2020).
6.3.3 Integrated Summary of Immunosuppression
Toxicological evidence for Pb-induced immunosuppression is derived from several lines of
evidence including functional assays (i.e., host resistance, antibody responses, DTH response, and ex vivo
WBC function) that are bolstered by various forms of supporting immune system data including
immunoglobulin levels, immune organ weight, WBC counts and differentials (immune organs), and WBC
counts (hematology). Toxicological studies evaluated in the 2013 Pb ISA (U.S. EPA, 2013) provide clear
evidence that host resistance to bacterial infection is compromised following Pb exposure. Evidence
available in 2013 also demonstrated that levels of antigen-specific IgM were unaffected in Pb-exposed
mice infected with Salmonella. However, levels of IgG2a were decreased and IgGl antibodies were
increased in these mice providing evidence for a shift toward Th2-type immune responses resulting in
decreased resistance to Salmonella. The potential for Pb exposure to result in immunosuppression was
further evaluated using the DTH assay. Based on a long history of use, the DTH assay is considered one of
the most predictive immunotoxicity tests available (Dietert et al„ 2010). Suppression of the DTH response
is a hallmark of Pb exposure and has been consistently reported in rats and chickens with PECO-relevant
BLLs as well as in other studies either not reporting BLLs or reporting BLLs outside the scope of this
ISA. The effects of Pb administration on the TDAR was also evaluated in the 2013 Pb ISA. Results from
these investigations were inconsistent with one study reporting a decrease in the antibody response (BLL
not reported) and another showing no effect in mice with high BLLs (i.e., 59-132 (.ig/dL). The effects of
Pb exposure on the functions of various WBCs under ex vivo conditions indicated that Pb exposure results
in (1) suppression of Thl-mediated immunity (i.e., suppressed Thl cytokine production [e.g., IFN-
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y] and DTH response); (2) altered macrophage function (e.g., increased ROS production, decreased NO
production); and (3) reduced monocyte/macrophage phagocytosis.
The 2013 Pb ISA also described toxicological evidence for effects of Pb exposure on various
supporting immune system endpoints (e.g., total serum immunoglobulins, immune organ weights, WBC
counts) that support data derived from immune function assays. Investigations of these endpoints are
limited in number, however, and due to differences in experimental design, challenging to interpret. For
example, inconsistent effects of Pb exposure on total serum IgE and IgG subtypes were described in the
2013 Pb ISA. Data reporting effects of Pb exposure on immune organ weight were limited to one study
reporting increased relative spleen weight and another study reporting decreased lymph node weight
following Pb exposure. Additional studies investigated the number and relative abundance of different
types of WBC in the spleen, thymus, lymph nodes and bone marrow following exposure to Pb, although
study design limitations and differences in the types of WBC assessed limit our ability to interpret these
data. In the only study reporting on WBC counts and subpopulation data collected in blood reviewed in
the 2013 Pb ISA, Pb exposure reduced the total number of WBC, lymphocytes, and monocytes.
The epidemiologic studies relevant to immunosuppression that were evaluated in the 2013 Pb
ISA (U.S. EPA, 2013) were more limited in number than the available toxicological evidence base.
Irrespective, these studies indicated some evidence of an association between BLLs and viral and
bacterial infections in children. None of the studies considered potential confounders, however, and most
analyzed populations with higher BLLs (means >10 (ig/dL). As described in the 2013 Pb ISA, some
epidemiologic studies also examined the effects of Pb exposure on WBC populations and cytokine levels.
Evaluation of these provided generally consistent evidence of inverse associations between BLLs and T
cell abundance in children, though most associations were seen with high concurrent BLLs (>10 (.ig/dL).
These results were coherent with the toxicological evidence base. Studies examining macrophages,
neutrophils, and NK cells and lymphocyte activation (i.e., HLA-DR expression) were largely
uninformative because of limitations associated with consideration of potential confounders and a lack of
information on concentration-response relationship.
Since the 2013 Pb ISA, there have been several epidemiologic studies published investigating
aspects of immunosuppression. Recent studies investigating associations between Pb exposure and
decreased host resistance examine populations with wider age-ranges and much lower mean and median
BLLs than studies evaluated in the 2013 Pb ISA. Recent studies also adjust for a wide range of potential
confounders, including extensive consideration of SES factors. Cross-sectional and case-control studies
provide consistent evidence of associations between Pb exposure and viral and bacterial infections or
susceptibility to antibiotic resistance. Antibody response, an endpoint that was not examined in studies
evaluated in the 2013 Pb ISA, was investigated in several recent studies. Specifically, a birth cohort study
and a few cross-sectional studies demonstrate generally consistent evidence of an association between
higher BLLs and lower counts of virus-neutralizing antibodies. A group of epidemiologic studies
examining children in China living either near an e-waste facility or in a nearby community with
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otherwise similar sociodemographic characteristics and pollutant exposures provides evidence that BLLs
are associated with differences in (1) the percentage of CD4+ naive and CD4+ central memory T cells, (2)
proinflammatory cytokine levels (IFN-y, IL-1|3, IL-8, IL-10, IL-12p70, and TNF-a), (3) levels of the
pleiotropic cytokine IL-6, (4) levels of the anti-inflammatory cytokine IL-10, and (5) the number of
neutrophils and monocytes. A few of the studies also reported null associations between BLLs and CD3+,
CD4+ and CD8+ cell counts, monocytes, and lymphocytes. The only recent study of an adult population
reported similar increases in cytokine levels associated with BLLs.
Available recent studies of immune function generally support evidence reported in the previous
Pb ISA. There are no recent toxicology studies investigating the effects of Pb exposure on host resistance
available for this review, but the strength of evidence reviewed in 2013 Pb IS As demonstrating that host
resistance to bacterial infection is compromised following Pb exposure has not diminished. A recent,
study shows, exposure to Pb had no effect on levels of anti-TT-specific IgM and IgG antibodies in rats.
However, levels of anti-TT-specific IgM (but not IgG) were decreased in iron-deficient rats. Consistent
with findings reported in the 2013 Pb ISA, recent studies also show that Pb exposures suppress the DTH
response, a widely-accepted measure of immunosuppression. Assessment of the effects of Pb exposure on
ex vivo WBC function is limited to assessments of Con A-stimulated lymphocyte proliferation and direct
measurement of cytokines in blood. Pb treatment had no effect on Con A-stimulated proliferation of
splenocytes collected from rats, however, treatment increased Con A-stimulated splenocyte proliferation
in iron-deficient rats. Pb exposure had no effect on levels of erythropoietin, GM-CSF, IL-6, and TNF-a in
a single study performed in rats. Recent studies reporting on the effects of Pb exposure on immune organ
pathology were inconsistent, with one study reporting effects on spleen architecture and another showing
no effect. Pb exposure reduced total serum IgA immunoglobulins in rats fed a control diet and in iron-
deficient rats but had no effect on total serum IgM and IgG in rats fed either diet. Recent investigations
also include assessments of the effects of Pb exposure on immune organ weight. Relative thymus weight,
which was not evaluated in the 2013 Pb ISA, decreased following exposure to Pb. As with the 2013 Pb
ISA, the effects of Pb exposure on relative spleen weight were inconsistent, varying with dose, exposure
duration, and route of administration (oral versus inhalation). Similarly, because of differences in
experimental design and the specific types of WBCs assessed in each study, it is difficult to interpret data
collected on the number and relative abundance of the different types of WBCs in the spleen, thymus,
lymph nodes and bone marrow following exposure to Pb. WBC counts and subpopulation data collected
from hematology investigations are similarly challenging to interpret.
6.4 Sensitization and Allergic Responses
Hypersensitivity responses are the result of an over-reaction of the immune system.
Hypersensitivity reactions are organized into four different classes, types I, II, III, and IV (Murphy and
Weaver. 2016). Irrespective of the type of response, all hypersensitivity responses develop in the same
two phases: sensitization and elicitation (or challenge). During the sensitization phase, the immune
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system is trained to respond to an otherwise innocuous antigen. This phase typically occurs without
symptoms. During the elicitation phase, the previously sensitized individual is re-exposed to the antigen
precipitating the symptoms of the allergic disease. Important for risk assessors, the concentration of the
sensitizing chemical required to elicit an allergic response is, in some cases, orders of magnitude lower
than the concentration required for sensitization. Consequently, preventing allergic sensitization from
developing in the first place is of paramount importance because dangerous, potentially life-threatening
allergic reactions can occur in response to exposure to a prohibitively-low concentration of the sensitizer.
6.4.1 Epidemiologic Studies of Sensitization and Allergic Responses
Epidemiologic studies of sensitization and allergic response generally cover studies of atopic
diseases, including asthma, rhinitis, and eczema, as well as studies examining cells and antibodies that
mediate these diseases, such as IgE and eosinophils. A limited number of studies evaluated in the 2013
Pb ISA (U.S. EPA, 2013) provide evidence of associations between exposure to Pb and asthma and
allergic sensitization. The strongest evidence comes from two prospective analyses, one investigating
incident asthma requiring medical care (Joseph et al„ 2005) and another examining allergic
hypersensitization via skin prick tests (SPTs) (Jedrychowski et al., 2011). Associations in both studies
were reported after adjustment for multiple confounders, including sex; birth weight; parity; maternal
age, education, and atopy; income; and prenatal and postnatal smoking exposure. Joseph et al. (2005)
observed associations between asthma incidence and BLLs >5 (ig/dL in white children (relative risk
[RR]: 2.7 [95% CI: 0.9, 8.1] compared with white children with BLL <5 (.ig/dL). In analyses restricted to
Black children, those with BLLs >10 (ig/dL had an elevated risk of incident asthma requiring medical
care (RR: 1.3 [95% CI: 0.6, 2.6] compared with children with BLLs <5 (.ig/dL). The effect estimates for
both groups were imprecise due to small numbers of children with asthma in the higher BLL categories
(five white children with BLLs >5 (ig/dL and nine Black children with BLLs >10 (.ig/dL). Jedrychowski et
al. (2011) also reported positive but imprecise associations (i.e., wide 95% CIs) between prenatal cord
BLLs and risk of positive SPT (rash/inflammatory reaction) to dust mite, dog, or cat allergen (RR: 2.3
[95% CI: 1.1, 4.6] for each 1 (ig/dL higher level of prenatal cord BLL). An additional prospective cohort
analysis reported an imprecise association between cord BLLs and prevalent asthma in children
(Rabinowitz et al„ 1990), but did not adjust for potential confounders and had low participation rates
with no information on nonparticipants. These findings were supported by a cross-sectional study of cord
blood and blood Pb-associated prevalent asthma (Pugh Smith and Nriagu, 2011). In addition to studies
examining atopic disease incidence or prevalence, the 2013 Pb ISA (U.S. EPA, 2013) also includes
supporting evidence from population-based cross-sectional studies in children that reported associations
between BLL and elevated serum IgE. Notably, many of these studies had limited adjustment for
potential confounders and included populations with mean BLLs >5 (ig/dL.
There have been several recent epidemiologic studies of sensitization and allergic response,
including prospective birth cohorts and cross-sectional studies with mean or median BLLs <2 (ig/dL. In
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general, these recent studies provide little evidence of an association between exposure to Pb and atopic
disease, and inconsistent evidence for immunological biomarkers involved in hypersensitivity and allergic
response. Measures of central tendency for BLL used in each study, along with other study-specific
details, including study population characteristics and select effect estimates, are highlighted in
Table 6-13. An overview of the recent evidence is provided below.
Whereas epidemiologic evidence from the 2013 Pb ISA supported the presence of an association
between BLL and incident and prevalent asthma in children, evidence from a few recent studies at lower
BLL is not indicative of an association. Specifically, in a small prospective birth cohort in France, Pesce
et al. (2021) reported that neither BLL measured during pregnancy nor cord BLL at birth were associated
with incident parental-reported asthma attacks through 5 years of age. Notably, there was a low rate of
asthma in the study population, limiting the statistical power to detect an association. However, because
asthma can be difficult to diagnose in children under 5, asthma attacks may be the most reliable measure.
The odds of asthma development associated with maternal BLLs were slightly elevated in the highest
quartile of exposure compared to the lowest, but the reported OR (1.25 [95% CI: 0.71, 2.2]) was
imprecise and the authors did not adjust the estimates for multiple comparisons (i.e., two exposure metrics
and four outcomes). In a cross-sectional NHANES analysis including slightly older children (2-12 years
old), Wells et al. (2014) also observed a null association between BLL and prevalent asthma.
Other recent epidemiologic studies of atopic disease are also generally consistent in reporting a
lack of an association with low levels of exposure to Pb. A few birth cohorts (Kim et al.. 2019; Kim et al..
2013) and a cross-sectional NHANES analysis including respondents of all ages (Wei et al.. 2019) did not
observe associations between cord blood or BLL and eczema incidence or prevalence. While Pesce et al.
(2021) reported a null association between maternal BLL and eczema in the aforementioned French birth
cohort, the authors did note substantially higher odds of eczema incidence for children in the higher
quartiles of cord blood Pb exposure compared with the lowest quartile. However, given the range of
outcomes examined (which included null associations for rhinitis and food allergy, in addition to asthma)
and the use of two exposure metrics (maternal blood and cord blood), the eczema results could be an
artifact of multiple testing. Consistent with Pesce et al. (2021). Mener et al. (2015) also reported a null
association between BLL and food allergies in children. However, the authors noted 10% higher odds of
food allergy sensitization in adults per 1 (ig/dL higher BLL (95% CI: 1%, 20%). In a restricted cubic
spline model, the observed relationship was approximately linear across the range of lower BLLs
(<3 (ig/dL), with no evidence of a threshold.
Results from a limited number of recent epidemiologic studies of allergen-specific and non-
specific immunological biomarkers of hypersensitivity in adults are inconsistent. A cross-sectional Korea
National Health and Nutrition Examination Survey (KNHANES) analysis reported higher total IgE
concentrations associated with higher BLLs in adults (Kim et al.. 2016). Notably, the observed
association was stronger in magnitude in respondents with house dust mite sensitization (10.4% [95% CI:
3.3%, 17.8%] per 1 (ig/dL higher BLL) compared with those without (3.5% [95% CI: -1.8%, 9.4%]). No
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other recent studies examined total IgE levels in adults, although Tsuii et al. (2019) reported that BLLs
were not associated, or slightly negatively associated, with allergen-specific serum IgE concentrations in
pregnant women, including egg white, hose dust mite, Japanese cedar pollen, animal dander, and moth
allergens. The interpretation of the results is complicated, however, by timing of the exposure and
outcome, where IgE concentrations were measured earlier in pregnancy (first trimester) than BLL (second
or third trimester).
Recent epidemiologic studies of non-specific immunological biomarkers of hypersensitivity in
neonates and children also provide inconsistent evidence of an association with exposure to Pb. In a small
birth cohort in south Korea, Kim et al. (2019) observed a cross-sectional association between higher cord
BLL and higher cord blood IL-13. In another cross-sectional analysis, Wells et al. (2014) reported that
each 1 (ig/dL higher level of BLL was associated with 10.3% (95% CI: 3.5%, 17.5%) higher serum total
IgE and 4.6% (95% CI: 2.4%, 6.8%) higher percent eosinophils. In contrast, results from a larger birth
cohort in Canada did not indicate higher odds of elevated cord blood IgE concentrations in relation to
higher average BLL across the first and third trimesters of pregnancy (Ashley-Martin et al.. 2015).
Further, the authors reported an inverse association between pregnancy BLL and odds of simultaneously
elevated cord blood IL-33 and thymic stromal lymphopoietin (TSLP).
6.4.2 Toxicological Studies of Sensitization and Allergic Responses
The 2013 Pb ISA reviewed evidence for the ability of Pb to induce immediate-type
hypersensitivity leading to the development of allergic asthma (U.S. EPA. 2013). Available studies
reported that exposure to Pb increased lymph node cell proliferation, increased production of Th2
cytokines such as IL-4, increased total serum IgE antibody levels in serum, and misregulated
inflammation. Recent toxicological evidence is limited in number and reports on the effects of Pb
exposure on production of cytokines relevant to immediate-type hypersensitivity, as discussed below.
6.4.2.1 Immediate-Type Hypersensitivity
Immediate-type hypersensitivity (i.e., type I) responses are the result of the production of IgE
antibodies, which trigger an array of responses, including anaphylaxis, allergic rhinitis, allergic
conjunctivitis, food allergy, atopic eczema, and allergic asthma. As with other forms of hypersensitivity,
immediate-type hypersensitivity develops in two stages. During the sensitization phase, antigen is
presented to naive T cells by antigen-presenting cells which promotes differentiation to the Th2
phenotype and the formation of memory T cells. Memory-specific T cells interact with antigen-specific B
cells leading the production of antigen-specific IgE antibodies that bind to Fc receptors on the surface of
mast cells. Upon secondary exposure to the allergen, the antigen binds to mast cell-bound IgE, triggering
mast cell degranulation resulting in eosinophil recruitment, mucus production, reactive airways and,
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potentially, anaphylaxis (Jancwav et al.. 2005). There are no validated animal models for determining
whether a xenobiotic can cause immediate-type hypersensitivity. For that reason, the potential for a
chemical to cause immediate-type hypersensitivity is assessed using a weight of the evidence approach
where data from an array of experimental endpoints (total serum IgE, antigen-specific IgE, eosinophilia of
the lung, measures of lung function, etc.) are carefully integrated (IPCS. 2012).
As reviewed in the 2013 Pb ISA, toxicological evidence, and to a lesser extent epidemiologic
evidence, have supported the effects of Pb exposure on stimulating Th2 activity. Studies have reported
increased lymph node cell proliferation (Teiion et al.. 2010; Carey et al.. 2006). increased production of
Th2 cytokines such as IL-4 (Fernandez-Cabezudo et al.. 2007; Iavicoli et al.. 2006; Chen et al.. 2004; Heo
et al.. 1998; Miller etal.. 1998; Heo et al.. 1997; Heo et al.. 1996). increased total serum IgE antibody
levels (Snyder et al.. 2000; Miller et al.. 1998; Heo et al.. 1997; Heo et al.. 1996). and misregulated
inflammation (Lodi et al.. 2011; Chettv et al.. 2005; Flohe et al.. 2002; Shabani and Rabbani. 2000; Miller
et al.. 1998; Chen et al.. 1997; Knowles and Donaldson. 1997; Bavkov et al.. 1996; Lee and Battles. 1994;
Zelikoff et al.. 1993; Knowles and Donaldson. 1990; Hilbertz et al.. 1986; Castranova et al.. 1980). These
endpoints comprise a well-recognized mode of action for the development and exacerbation of atopic and
inflammatory conditions such as asthma and allergy.
Only two recent toxicology studies investigated the effects of Pb exposure on production of
cytokines relevant to immediate-type hypersensitivity. In one of these studies, administration of Pb
acetate drinking water (300 ppm; BLL = 18.48 (ig/dL) for 42 days decreased IFN-y levels, but had no
effect on IL-10 levels (data not shown) in adult male Sprague Dawley rats (Fang et al.. 2012). In addition,
administration of Pb acetate in drinking water (0.2%; BLL = 9.3 (ig/dL) for 84 days had no effect on
erythropoietin, GM-CSF, IL-6, and TNF-a levels in blood collected from Sprague Dawley rats (data from
sexes pooled) (Cai et al.. 2018). Study-specific details, including animal species, strain, sex and BLLs, are
highlighted in Table 6-14.
6.4.3 Integrated Summary of Sensitization and Allergic Responses
As reviewed in the 2013 Pb ISA (U.S. EPA, 2013), toxicological evidence, and to a lesser extent
epidemiologic evidence, have supported the effects of Pb exposure on increased lymph node cell
proliferation, increased production of Th2 cytokines such as IL-4, increased total serum IgE antibody
levels in serum, and misregulated inflammation. Additionally, a limited number of longitudinal
epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA, 2013) provide evidence of associations
between exposure to Pb and asthma (Joseph et al., 2005) and allergic sensitization (Jedrychowski et al..
2011). The associations in these studies are imprecise (i.e., wide 95% CIs), but are supported by cross-
sectional studies of cord blood and blood Pb-associated prevalent asthma and population-based cross-
sectional studies in children that reported associations between BLL and elevated serum IgE (U.S. EPA,
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2013). Many of these cross-sectional studies had limited adjustment for potential confounders and
included populations with mean BLLs >5 (ig/dL.
Though limited in number, recent PECOS-relevant animal toxicological studies continue to
support the findings from the last review. Specifically, these studies consistently report effects of Pb on
sensitization and allergic responses including two studies of the effects of Pb exposure on production of
cytokines relevant to immediate-type hypersensitivity. In contrast, recent epidemiologic evidence is not
consistent with studies evaluated in the 2013 Pb ISA. Specifically, recent studies provide little evidence
of an association between exposure to Pb and atopic disease, and inconsistent evidence for
immunological biomarkers involved in hypersensitivity and allergic response. Similar to cohort studies
evaluated in the 2013 Pb ISA, recent longitudinal analyses are limited in number and have limited
statistical power because of small case numbers. Limited statistical power results in the reduced
likelihood of detecting a true effect and a reduced likelihood that an observed result reflects a true effect.
Whereas there was coherence between the animal toxicological and epidemiologic evidence evaluated in
the 2013 Pb ISA, the recent evidence is less coherent given the inconsistencies and null findings across
epidemiologic studies.
6.5 Autoimmunity and Autoimmune Disease
Autoimmunity is characterized by the reaction of autoreactive T lymphocytes or autoantibodies
against self-molecules (i.e., autoantigens). Depending on the etiology, autoimmunity may lead to the
development of autoimmune diseases such as rheumatoid arthritis and lupus. While the precipitating
event for the development of autoimmunity is often unknown, intrinsic factors (e.g., gene polymorphisms,
sex-related hormones, and age) and extrinsic factors (e.g., lifestyle, exposure to certain drugs, chemicals,
and infectious agents) are known to play a role in the induction, development, or exacerbation of
autoimmunity (IPCS. 2012). Although animal models have been used to study a variety of autoimmune
diseases, there are currently no validated models to assess or identify chemicals that induce or exacerbate
autoimmune diseases (IPCS. 2012). Consequently, the potential to induce or exacerbate autoimmunity is
best investigated using atiered approach composed of multiple methods. The 2013 Pb ISA concluded the
available toxicological and epidemiologic studies were inadequate to infer that a causal relationship exists
between Pb exposure and the development of autoimmunity and autoimmune disease.
6.5.1 Epidemiologic Studies of Autoimmunity and Autoimmune Disease
A single epidemiologic study evaluated in the 2013 Pb ISA (U.S. EPA. 2013) examined the
association between exposure to Pb and autoimmunity (El-Fawal et al.. 1999). While the authors reported
higher levels of autoantibodies in Pb-exposed battery workers, the analysis did not include adjustment for
important confounders (e.g., other occupational exposures) and included BLLs of 10-40 (ig/dL, much
higher than those found in the general population. Recent epidemiologic studies of autoimmunity are
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limited in number and examine disparate outcomes. Mean BLL used in each study, along with other
study-specific details, including study population characteristics and select effect estimates, are
highlighted in Table 6-15. An overview of the recent evidence is provided below.
Two recent population-based cross-sectional studies provide inconsistent evidence of associations
between exposure to Pb and autoimmune disorders (Joo et al.. 2019; Kamvcheva et al.. 2017). In an
NHANES analysis of seropositivity for Celiac Disease (i.e., tissue transglutaminase [tTg]-IgA),
Kamvcheva et al. (2017) reported lower adjusted mean BLLs in children with Celiac Disease compared
with those without (-0.14 (ig/dL [95% CI: -0.27, -0.02 |ig/dL|). Associations were comparable in
magnitude, but less precise in adults (i.e., wider 95% CIs). Cross-sectional studies cannot establish
temporality and the nature of malabsorption in Celiac Disease makes it biologically plausible that the
disorder could result in reduced absorption of Pb rather than there being a protective effect of Pb
exposure. Another population-based study did not observe an association between BLL and rheumatoid
arthritis (Joo et al.. 2019). A notable limitation of this study is that it included children, while rheumatoid
arthritis primarily affects adults.
6.5.2 Toxicological Studies of Autoimmunity and Autoimmune Disease
As reported in the 2013 Pb ISA, evidence for the ability of Pb to induce autoimmunity is limited
(U.S. EPA. 2013). Only one study performed in rats showed the generation of autoantibodies following
Pb administration by a relevant route of exposure (i.e., dietary) (El-Fawal et al.. 1999). Several other
studies utilized Pb exposure routes or doses that produced BLLs that are not relevant to humans (Hudson
et al.. 2003; Bunn et al.. 2000; Waterman et al.. 1994). There is only one recent toxicology study that
investigates an endpoint directly related to the development of autoimmunity. In that study, Fang et al.
(2012) reported that administration of Pb acetate in drinking water for 42 days (BLL = 18.48 (ig/dL) had
no effect on the suppressive properties of Tregs isolated from adult male Sprague Dawley rats. Study-
specific details, including animal species, strain, sex, and BLLs are highlighted in Table 6-16.
6.5.3 Integrated Summary of Autoimmunity and Autoimmune Disease
An epidemiologic study evaluated in the 2013 Pb ISA (U.S. EPA. 2013) observed an association
between higher BLLs and elevated autoantibodies, but the strength of conclusions that can be drawn from
this study is limited because it did not control for important confounders. Toxicological evidence
demonstrating that Pb exposure leads to autoimmunity is similarly limited. As discussed in the 2013 Pb
ISA (U.S. EPA. 2013). one PECOS-relevant study and several other studies utilizing non-PECOS routes
of exposure and doses that produced BLLs that are not relevant to humans showed the generation of
autoantibodies following Pb administration. Recent epidemiologic studies of autoimmunity are limited in
number, examine disparate outcomes and provide inconsistent evidence of associations between exposure
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to Pb and autoimmune disorders. A recent toxicological study reported that Pb exposure had no effect on
the suppressive properties of Tregs, which are critical mediators of immune tolerance.
6.6 Biological Plausibility
This section describes biological pathways that potentially underlie effects on the function of the
immune system resulting from exposure to Pb. Figure 6-1 depicts the proposed pathways as a continuum
of upstream events, connected by arrows, that may lead to downstream events observed in epidemiologic
studies. Evidence supporting these proposed pathways was derived from Sections 6.3, 6.4, and 6.5 of
this ISA, evidence reviewed in the 2013 Pb ISA (U.S. EPA, 2013), and recent evidence collected from
studies that may not meet the current PECOS criteria, but contain mechanistic information supporting
these pathways. This discussion of how exposure to Pb may lead to immune system effects contributes
to an understanding of the biological plausibility of epidemiologic results evaluated later in the ensuing
sections. Note that the structure of the Biological Plausibility section and the role of biological
plausibility in contributing to the weight-of-evidence analysis used in the 2013 Pb ISA are discussed
below.
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Altered dendric
cell function
Increase Th2
cytokines
{e.g., IL-4)
Antibody
production/secretion
Class switching
Increased IgE secretion
Allergic asthma
Pb
Exposure
h
Increase proinflammatory
cytokines
(e.g.,TNF-a)
Increase ROS production
Increase cell death
Decrease phagocyte function
Decrease chemotaxis
function
Decrease nitric oxide
production
Immunosuppression/
increased incidence of
infection
DTH = delayed-type hypersensitivity; IgE = immunoglobulin E; IFN-y = interferon-gamma; IL-4 = interleukin 4; ROS = reactive
oxygen species; Th2 = T helper; TNF-a = tumor necrosis factor alpha.
Note: The boxes above represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and
the arrows indicate a proposed relationship between those effects. Solid arrows denote evidence of essentiality as provided, for
example, by an inhibitor of the pathway used in an experimental study involving Pb exposure. Dotted arrows denote a possible
relationship between effects. Shading around multiple boxes is used to denote a grouping of these effects. Arrows may connect
individual boxes, groupings of boxes, and individual boxes within groupings of boxes. Progression of effects is generally depicted
from left to right and color coded (white, exposure; green, initial effect; blue, intermediate effect; orange, effect at the population
level or a key clinical effect). Here, population-level effects generally reflect results of epidemiologic studies. When there are gaps in
the evidence, there are complementary gaps in the figure and the accompanying text below. The structure of the biological
plausibility sections and the role of biological plausibility in contributing to the weight-of-evidence analysis used in the 2024 Pb ISA
are discussed in Section 6.7.
Figure 6-1 Potential biological plausibility pathways for immunological
effects associated with exposure to Pb.
Immunotoxicity may be expressed as immunosuppression, unintended stimulation of immune
responses, hypersensitivity, or autoimmunity (IPCS. 2012). The World Health Organization's Guidance
for Immunotoxicity Risk Assessment for Chemicals (IPCS. 2012) describes best approaches for weighing
immunotoxicological data. Within this framework, data from endpoints observed in the presence of
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immune challenge (e.g., including effects on antibody responses, host resistance, and ex vivo WBC
function) are considered most informative whereas other measures collected in the absence of immune
stimulation (e.g., immune organ pathology, non-specific immunoglobulin levels, WBC counts,
lymphocyte subpopulations, T cell subpopulations, immune organ weights) are considered supporting
evidence. Careful review of the evidence base suggests that exposure to Pb has the potential to modulate
the immune system leading to immunosuppression and sensitization and allergic responses. Below,
evidence from peer-reviewed toxicology studies providing biological plausibility for Pb-associated
immunotoxicity is reviewed.
6.6.1 Immunosuppression
Immunosuppression can lead to the increased incidence and severity of infectious and neoplastic
diseases. Importantly, there are internationally validated animal models and human correlates (e.g., the
rodent DTH assay and the human tuberculin test) for assessing the potential for a chemical to induce
immunosuppression. Still, the potential for a chemical to suppress the function of the immune system is
best assessed using a weight of the evidence approach where data from an array of experimental
endpoints are carefully integrated (IPCS. 2012).
The initiating event that ultimately leads to Pb-induced immunosuppression is unknown.
However, Pb exposure has been shown to affect several indicators of immunosuppression including
decreased Thl cytokine production, production of other inflammatory mediators, decreased macrophage
function (chemotaxis and phagocytosis), and ultimately suppressed the DTH response (Figure 6-1).
Exposure to Pb has been convincingly shown to result in the skewing of T cell populations,
simultaneously promoting the formation of Th2 cells while suppressing the formation of Thl cells and
their cytokines including IFN-y that play key roles in cell-mediated immunity (Heo et al.. 1996; Fochtman
et al.. 1969). Available evidence suggests that this phenomenon may involve Pb-induced effects on
dendritic cells, which promote skewing towards the Th2 phenotype (Gao et al.. 2007). Mitogen-
stimulated production of IFN-y was significantly lower in splenocytes collected from Pb-exposed mice
(Dvoroznakova and Jalcova. 2013). IFN-y levels in serum were reduced in Pb-exposed mice (Aiouaoi et
al.. 2020). IFN-y is the primary cytokine that stimulates recruitment of macrophages associated to sites of
inflammation (Lee et al.. 2001; Chen et al.. 1999). Relevant decrements in macrophage function
associated with Pb exposure have been reported, including decreased chemotaxis (Lodi et al.. 2011;
Bishavi and Sengupta. 2006) and phagocytosis (Lodi et al.. 2011; Bussolaro et al.. 2008; Bishavi and
Scngupta. 2006; Hilbertz et al.. 1986; Zhou et al.. 1985; Castranova et al.. 1980). Macrophages play a
vital role in cell-mediated immunity, which is often assessed using the DTH response when assaying
potential immunosuppressants. Pb exposure has been consistently shown to suppress the DTH response in
rodents with BLLs relevant to human exposures. Observations of a concomitant decrease in IFN-y
strengthen the link between Pb-induced inhibition of Thl functional activities and suppression of the
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DTH response (Lee et al.. 2001; Chen et al.. 1999). Furthermore, the effects of Pb exposure on
macrophage PGE2 (Chettv et al.. 2005). decreased ROS production (Chen et al.. 1997; Hilbertz et al..
1986; Castranova et al.. 1980). decreased NO production (Farrer et al.. 2008; Mishra et al.. 2006; Bunn et
al.. 2001b; Lee et al.. 2001; Krocova et al.. 2000; Chen et al.. 1997; Tian and Lawrence. 1996; Tian and
Lawrence. 1995). and increased cell death (Metrvka et al.. 2021; Guan et al.. 2020; Choi et al.. 2018; Kerr
et al.. 2013) may contribute to decreased resistance to bacterial or viral infection (Hilbertz et al.. 1986;
Castranova et al.. 1980). Pb exposure has also been shown to increase levels of TNF-a, a
proinflammatory cytokine, secreted by LPS- stimulated mouse J774A. 1 macrophages (Luna et al.. 2012)
and human THP-1 monocytes through a mechanism involving ERK1/2 (Khan et al.. 2011). As reviewed
in the 2006 Pb AQCD (U.S. EPA. 2006). Pb exposure also has the potential to reduce neutrophil
chemotaxis, phagocytosis, and respiratory oxidative burst, but the effect was not judged to be as strong as
what has been observed in relation to macrophages. Finally, decreased Thl signaling leading to
differences in IgG isotypes produced in response to S. enterica infection was implicated in impaired host
defense in mice (Fernandez-Cabezudo et al.. 2007).
While there is compelling evidence that Pb exposure can decrease host resistance to infection, the
effect may not be attributable to direct effects of Pb exposure on the immune system. Instead, decreased
host resistance may be the result of Pb acting on the microbiome. The microbiome is the body's gateway,
disruption of microbiome can have profound effects on xenobiotic processing, and resistance to pathogens
(Zhai et al.. 2020; Dietert and Silbergeld. 2015; Nriagu and Skaar. 2015). The human microbiome
comprises most of the cells and genes in the human body, and these cells are the first to be exposed to
environmental chemicals. The microbiome plays a key role in excretion levels, transport barriers
(e.g., skin, lung, gut barriers), metabolism of xenobiotics (Zhai et al.. 2020; Dietert. 2018; Nriagu and
Skaar. 2015). In addition, changes in the composition of the microbiome following exposure to
xenobiotics can affect the process of colonization resistance to pathogens which may lead to loss of
mucosal barrier function, elevated risk of infection, and the development of noncommunicable diseases
such as asthma (Huang et al.. 2020; Zhai et al.. 2020; Dietert. 2018; Nriagu and Skaar. 2015).
Importantly, Pb is known to possess antimicrobial properties(Mivano et al.. 2007). As reviewed by Liu et
al. (2021). exposure to Pb has been shown to alter the diversity and relative composition of the gut
microbiota in several toxicology studies performed in laboratory animals. Our ability to interpret these
findings is limited, however, by the fact that the investigators conducting these studies either did not
measure BLL at all or, in the two studies that did, the BLL was not relevant to human exposure. In
addition to toxicological studies, a limited number of epidemiologic studies reported associations between
biomarkers of Pb exposure and altered gut microbiota diversity, including a birth cohort study (Sitarik et
al.. 2020) and a few cross-sectional analyses (Zeng et al.. 2022; Eggers et al.. 2019). Further, the
possibility that the effects of Pb on the immune system are at least partly mediated by the microbiome is
supported by the capacity of certain probiotics to protect against Pb-induced toxicity (i.e., decreases BLL
and relieves Pb-induced intestinal barrier impairment) in mice (Zhai et al.. 2020). In rats, chelation
treatment reduced IL-4 production and IFN-y suppression induced by Pb (Chen et al.. 1999). Similarly,
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Vitamin D supplementation was shown to reduce Pb-induced IL-4 in rats, but the concentration of IL-4
remained significantly elevated relative to control (BaSalamah et al.. 2018).
6.6.2 Sensitization and Allergic Responses
Hypersensitivity responses (i.e., allergies) are the result of an over-reaction of the immune
system. Immediate-type hypersensitivity responses are the result of the production of IgE antibodies,
which trigger an array of responses including anaphylaxis, allergic rhinitis, allergic conjunctivitis, food
allergy, atopic eczema, and allergic asthma. Like with other forms of hypersensitivity, immediate-type
hypersensitivity, develops in two stages. During the sensitization phase, antigen is presented to naive T
cells by antigen-presenting cells, which promotes differentiation to the Th2 phenotype and the formation
of memory T cells. Memory-specific T cells interact with antigen-specific B cells leading the production
of antigen-specific IgE antibodies that bind to Fc receptors on the surface of mast cells. Upon secondary
exposure to the allergen, the antigen binds to mast cell-bound IgE, triggering mast cell degranulation
resulting in eosinophil recruitment, mucus production, reactive airways and, potentially, anaphylaxis
(Janewav et al.. 2005). Importantly, there are no validated animal models for determining whether a
xenobiotic can cause allergic asthma. For that reason, the potential for a chemical to cause allergic asthma
is assessed using a weight of the evidence approach where data from an array of experimental endpoints
are carefully integrated (IPCS. 2012).
The initiating event that ultimately leads to allergic sensitization is called haptenation, the process
where sensitizing chemical binds to endogenous proteins leading to detection by the immune system and
ultimately allergic sensitization (Janewav et al.. 2005). To date, there are no publications demonstrating
that Pb acts as a hapten. Pb exposure, however, is associated with other hallmarks of allergic
hypersensitivity and asthma including Th2 cytokine production, B cell activation, and production of IgE
antibodies that are central to these responses.
Exposure to Pb resulting in BLLs relevant to humans has been convincingly shown to result in
the skewing of T cell populations, simultaneously suppressing the formation of Thl cells while promoting
the formation of Th2 cells and cytokines that promote the development of allergic airway disease (Heo et
al.. 1996; Fochtman et al.. 1969). IL-4 is a key regulator of immune responses produced by Th2 cells.
This pleiotropic cytokine not only inhibits production of Thl cytokines, but also promotes B cell
activation, differentiation, proliferation and class switching leading to the production of IgE antibodies
(Dietert and Piepenbrink. 2006). Importantly, in most cases where Pb exposure was associated with
increased IgE levels, IL-4 levels were also elevated (Snyder et al.. 2000; Chen et al.. 1999; Miller et al..
1998). IgE antibodies are a hallmark of immediate-type hypersensitivity responses that are responsible for
inducing allergic asthma (Janewav et al.. 2005). In sensitized individuals, binding of allergen to antigen-
specific IgE antibodies on the surface of mast cells triggers mast cell degranulation and release histamine,
leukotrienes, and cytokines, which in turn, produce the inflammatory-related effects associated with
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asthma and allergy, i.e., airway responsiveness, mucus secretion, respiratory symptoms (Jancwav et al..
2005). Consistent with this condition, inflammation was identified as a major immune-related effect of Pb
based on consistent toxicological evidence for Pb-induced increases in proinflammatory cytokines
(e.g., IL-4) and increased levels of PGE2 (Chettv et al.. 2005) and ROS production (Chen et al.. 1997;
Hilbertz et al.. 1986; Castranova et al.. 1980). decreased NO production (Farrer et al.. 2008; Mishra et al..
2006; Bunn et al.. 2001b; Lee et al.. 2001; Krocova et al.. 2000; Chen et al.. 1997; Tian and Lawrence.
1996; Tian and Lawrence. 1995). and increased cell death (Metrvka et al.. 2021; Guan et al.. 2020; Choi
et al.. 2018; Kerr et al.. 2013) that may also contribute to Pb-induced decreased resistance to bacterial or
viral infection (Hilbertz et al.. 1986; Castranova et al.. 1980).
6.7 Summary and Causality Determinations
The body of epidemiologic and toxicological evidence describes several effects of Pb exposure
on the immune system. The majority of this evidence predates this ISA. These effects can be traced back
to two major targets including T cells and macrophages promoting immunosuppression and sensitization
and allergic responses, respectively. In addition, a very limited number of studies report findings related
to autoimmunity. The sections that follow describe the evaluation of evidence for these three groups of
outcomes with respect to causality determinations for exposure to Pb using the framework described in
the Preamble to the ISA (U.S. EPA. 2015). The key evidence, as it relates to the causal framework, is
outlined below, and summarized in Table 6-1, Table 6-2, and Table 6-3.
6.7.1 Causality Determination for Immunosuppression
The 2013 Pb ISA concluded "a causal relationship is likely to exist between Pb exposures and
decreased host resistance."(U.S. EPA, 2013). This causality determination was primarily based on
consistent evidence that exposure to relevant BLLs suppresses the DTH response and increases bacterial
titers and subsequent mortality in rodents. For example, various studies reported decreased clearance of
bacteria and increased mortality induced by Listeria monocytogenes in mice exposed postnatally to Pb
acetate in drinking water for 3 to 8 weeks, resulting in BLL ranging from 20-25 (ig/dL (Fernandez-
Cabezudo et al., 2007; Dyatlov and Lawrence, 2002; Kim and Lawrence, 2000; Kishikawa et al., 1997;
Lawrence, 1981). Other studies reported increased mortality from Salmonella or E. coli, or decreased
clearance of Staphylococcus, in mice administered Pb acetate or Pb nitrate via injection resulting in BLL
relevant to the 2013 Pb ISA (Bishayi and Sengupta, 2006; Cook et al.. 1975; Hemphill et al.. 1971; Selye
et al., 1966). Although BLLs were high (i.e., 71-313 (.ig/dL). increased mortality from viral infection was
also reported in mice and chickens administered Pb (mostly Pb acetate) for 4-
10 weeks (Gupta et al., 2002; Exon et al., 1979; Thind and Khan, 1978). Additional evidence for Pb-
induced immunosuppression comes from studies investigating the DTH response. Suppressed DTH
response is one of the most consistently reported immune effects associated with Pb exposure in animals
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(U.S. EPA. 2013). Suppression of the DTH response has been reported following gestational (Chen et al..
2004; Bunn et al., 2001a; Bunn et al.. 2001b, c; Lee et al., 2001; Chen et al., 1999; Miller et al., 1998;
Faith et al.. 1979) and postnatal (McCabe et al., 1999; Haneef et al., 1995; Laschi-Loquerie et al., 1984;
Miilleretal., 1977) exposures to Pb acetate resulting in BLLs ranging from 6.75 to >100 (ig/dL) in rats,
mice, chickens, and goats. Further, evidence suggested a plausible mode of action involving suppressed
production of Thl cytokines (e.g., IFN-y) (Fernandez-Cabezudo et al„ 2007; Lara-Tcjcro and Panicr.
2004), and decreased macrophage function (Lodi et al„ 2011; Bishayi and Sengupta, 2006; Chen et al„
1997; Hilbertz et al., 1986; Castranova et al., 1980). A limited number of epidemiologic studies reviewed
in the 2013 Pb ISA (U.S. EPA, 2013) indicated an association between BLL and viral and bacterial
infections in children. None of the studies considered potential confounders, however, and most analyzed
populations with higher BLLs (means >10 (ig/dL). Cross-sectional studies of cell-mediated immunity
reported consistent associations between BLL and lower T cell abundance in children, while results from
other studies on lymphocyte activation, macrophages, neutrophils, and NK cells were generally
inconsistent or not sufficiently informative (e.g., cross-sectional study designs with limited or no
consideration of potential confounding, and a lack of information on concentration-response relationship).
Recent toxicological studies provide additional evidence for immunosuppression. Although there
were no recent studies directly investigating the effects of Pb exposure on host resistance, the ability of Pb
to alter antibody responses was investigated and provides evidence for immunosuppression. Yathapu et
al. (2020) showed that serum levels of anti-TT specific IgM antibodies were decreased while anti-TT
specific IgG levels were unaffected in rats exposed to Pb (BLL = 16.1 (ig/dL) in drinking water.
Consistent with the 2013 Pb ISA, administration of Pb acetate in drinking water for 42 days
(BLL = 18.48 (ig/dL) significantly suppressed the DTH response in adult male Sprague Dawley rats
(Fang et al., 2012). Additional supporting evidence for Pb-induced immunosuppression can be derived
from supporting immune system endpoints including (1) reduced non-specific mucosal IgA
immunoglobulins (but not IgM or IgG) in rats with BLLs of 16.1 (ig/dL (Yathapu et al., 2020) and (2)
reduced relative thymus weight in juvenile rats orally administered Pb (1 or 10 mg/kg with BLL of
3.27 (ig/dL and 12.5 (ig/dL, respectively) for up to 25 days (Graham et al., 2011). Because of differences
in experimental design parameters and specific endpoints measured, effects of Pb exposure on immune
organ pathology, WBC counts and differentials, and WBC counts (hematology and subpopulations) are
challenging to interpret and, for that reason, do not support or refute evidence obtained from immune
function assays.
The relationship between Pb exposure and immunosuppression is further supported by recent
epidemiologic studies, which expand quantity and quality of the supporting immune system evidence
base evaluated in the 2013 Pb ISA. Recent case-control and cross-sectional studies provide consistent
evidence that BLLs are associated with greater susceptibility to viral and bacterial infection in children
and adults (Feiler et al., 2020; Park et al., 2020; Krucgcr and Wade, 2016) and reduced antibiotic
resistance in children, as measured by nasal Staphylococcus aureus colonization (Eggers et al„ 2018).
Associations were observed with mean, median, or geometric mean BLLs <3.5 (ig/dL. The evaluated
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studies used concurrent blood Pb measures, raising uncertainty regarding the temporal sequence between
Pb exposure and immunosuppression and the magnitude, timing, frequency, and duration of Pb exposures
that contributed to the observed associations. Recent studies also provide generally consistent evidence of
an inverse association between BLLs and vaccine antibodies in children with low mean or median BLLs,
including a birth cohort of vaccinated children in South Africa with median BLLs <2 (ig/dL Di Lenardo et
al. (2020). A strength of this analysis is that it establishes temporality between exposure and outcome.
Cross-sectional studies, including a large analysis of children ages 6 to 17 from the 1990-2004 NHANES
(Jusko et al.. 2019). are consistent with results from the prospective birth cohort. Notably, this study
includes many children who were born before the phaseout of leaded gasoline and were likely subject to
higher past exposures. Thus, there is uncertainty concerning the specific Pb exposure level, timing,
frequency, and duration contributing to the associations observed in this study.
In summary, the collective body of evidence indicates that there is likely to be a causal
relationship between Pb exposure and immunosuppression. The strongest evidence supporting a
'likely to be causal" relationship between Pb exposure and immunosuppression comes from toxicological
studies consistently demonstrating that Pb exposures suppress the DTH response and increase
susceptibility to bacterial infection in animals with BLLs < 30 |ag/dL. These toxicological studies are
coherent with recent case-control and cross-sectional epidemiologic studies providing consistent evidence
that higher BLLs are associated with greater susceptibility to viral and bacterial infection in children and
adults and lower antibiotic resistance in children. Though these epidemiologic studies used concurrent
blood Pb measures, raising uncertainty regarding the temporal sequence between Pb exposure and
immunosuppression, a smaller body of supporting epidemiologic studies provide evidence that prenatal
(mean < 4 (.ig/dL). as well as concurrent (mean and/or medians < 2 (.ig/dL). BLLs are associated with a
smaller vaccine antibody response. The two toxicological studies examining the animal correlate for the
human vaccine response (i.e., TDARto sheep red blood cells) reported mixed results. The biological
plausibility of Pb-induced immunosuppression is supported by toxicological studies demonstrating (1)
skewing of T cell populations, promoting Th2 cell formation and cytokine production, (2) decreased IFN-
y production, (3) decrements in macrophage function, (4) production of inflammatory mediators, and (5)
disruption of the microbiome.
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Table 6-1 Summary of evidence for a likely to be causal relationship between Pb exposure and
immunosuppression
Rationale for Causality
Determination3
Key Evidence"
Key References"
Pb Biomarker Levels Associated with
Effects0
Consistent evidence from
toxicological studies with
relevant exposures
investigating immune
functional endpoints
Oral Pb exposures increased bacterial
infection. Similar observations in several
other studies using non-PECOS routes of
exposure and/or higher Pb exposures
Dvatlov and Lawrence (2002)
Fernandez-Cabezudo et al. (2007)
Mean BLL:
20 |jg/dL after adult 16-wk exposure
25 |jg/dL after lactational exposure
Oral gestational Pb exposures
suppressed DTH response. Similar
observations in several other studies with
higher Pb exposures
Chen et al. (2004)
Bunn et al. (2001a)
Fang et al. (2012)
Mean BLL:
6.75 |jg/dL
25 |jg/dL
18.48 |jg/dL
Evidence from other
toxicological studies with
relevant exposures
investigating immune
functional endpoints
Oral Pb exposure decreased levels of
anti-TT-specific IgM, levels of anti-TT-
specific IgG were unaffected
Yathapu et al. (2020)
Mean BLL:
16.1 ± 5.5 |jg/dL
Supporting evidence from
toxicological studies with
relevant exposures supporting
immune functional endpoints
Oral Pb exposure decreased non-specific
mucosal IgA immunoglobulins
Oral administration of Pb decreased
relative thymus weight in juvenile rats
Yathapu et al. (2020)
Graham et al. (2011)
Mean BLL:
16.1 ± 5.5 |jg/dL
1 or 10 mg/kg exposure dose with BLL of
3.27 |jg/dL and 12.5 |jg/dL, respectively
Coherence from a small body
of epidemiologic studies
demonstrating consistent
evidence of decreased host
resistance at low BLLs
A limited number of case-control and
cross-sectional studies reported
associations between concurrent BLLs
and:
Increased susceptibility to viral and
bacterial infection, and
Krueaer and Wade (2016)
Park et al. (2020)
Feiler et al. (2020)
Mean, Median, or Geometric Mean BLL
across studies:
1.4-3.15 |jg/dL
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Reduced antibiotic resistance
Eqqers et al. (2018)
Uncertainty regarding the temporal
sequence between Pb exposure and
immunosuppression and the magnitude,
timing, frequency, and duration of Pb
exposures that contributed to the
observed associations.
Coherence from a small body
of epidemiologic studies
demonstrating consistent
evidence of decreased
vaccine antibody response at
low BLLs
Jusko et al. (2019) Mean BLL: 1.4 pg/dL
See Section 6.3.1.2
A limited number of prospective birth Pi Lenardo et al. (2020) Median BLL: 1.9 pg/dL
cohort and cross-sectional studies
reported associations between BLLs and
decreased vaccine antibody response
Biological Plausibility Evidence that Pb (1) suppressed See Section 6.6
production of Th1 cytokines (i.e., IFN-y),
(2) decreased macrophage function, and
(3) increased inflammation in animals
anti-TT = anti-tetanus toxoid; BLL = blood lead level; DTH = delayed-type hypersensitivity; IgG = immunoglobulin G; IgM = immunoglobulin M; Pb = lead; PECOS = population,
exposure, comparator, outcome and study.
"Based on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
'Describes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
Describes the Pb biomarker levels at which the evidence is substantiated.
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6.7.2
Causality Determination for Sensitization and Allergic Responses
The 2013 Pb ISA concluded "that a causal relationship is likely to exist between Pb exposures
and an increase in atopic and inflammatory conditions" (U.S. EPA. 2013). This causality determination
was made on the basis of a body of evidence integrated across epidemiologic and toxicological studies.
Epidemiologic evidence included a prospective analysis reporting associations between BLLs and asthma
incidence in children (Joseph et al., 2005) and another longitudinal study that observed an association
between cord BLLs and immediate-type allergic responses in children that were detected clinically using
SPTs (Jedrychowski et al.. 2011). Both studies had small sample sizes, however, and lacked precision
(i.e., had wide 95% CIs), which increases the likelihood of chance findings. An additional prospective
cohort analysis reported an imprecise association between cord BLLs and prevalent asthma in children
(Rabinowitz et al., 1990) but did not adjust for potential confounders. The associations observed in the
prospective analyses are supported by a cross-sectional study of BLL-associated parental-reported asthma
in children and population-based cross-sectional studies in children that reported associations between
BLLs and elevated serum IgE. Notably, many of the serum IgE studies had limited adjustment for
potential confounders and included population mean BLLs >5 (ig/dL. The epidemiologic findings are
coherent with a large body of toxicological studies that reported physiological responses in animals
consistent with the development of allergic sensitization, including increased lymph node cell
proliferation (Teijon et al„ 2010; Carey et al„ 2006), increased production of Th2 cytokines such as IL-4
(Fernandez-Cabezudo et al., 2007; Iavicoli et al., 2006; Chen et al., 2004; Heo et al., 1998; Miller et al.,
1998; Heo et al„ 1997; Heo et al., 1996), increased total serum IgE antibody levels (Snyder et al., 2000;
Miller et al., 1998; Heo et al., 1997; Heo et al., 1996), and misregulated inflammation (Lodi et al., 2011;
Chetty et al„ 2005; Flohe et al., 2002; Shabani and Rabbani, 2000; Miller etal., 1998; Chen et al„ 1997;
Knowles and Donaldson, 1997; Baykov et al., 1996; Lee and Battles, 1994; Zelikoff et al., 1993; Knowles
and Donaldson, 1990; Hilbertz et al., 1986; Castranova et al., 1980).
There have been several recent epidemiologic studies of sensitization and allergic response,
including prospective birth cohorts and cross-sectional studies with mean or median BLLs <2 (ig/dL. In
contrast to evidence presented in the 2013 Pb ISA (U.S. EPA. 2013). the recent studies provide little
evidence of an association between exposure to Pb and atopic disease, and inconsistent evidence for
immunological biomarkers involved in sensitization and allergic response. Specifically, recent
epidemiologic studies of atopic disease, including analyses of prospective cohort studies examining of
asthma (Pesce et al., 2021), eczema (Pesce et al., 2021; Kim et al., 2019; Kim et al„ 2013), and food
allergies (Pesce et al„ 2021) were generally consistent in reporting a lack of an association in populations
with low mean BLLs. A considerable uncertainty in the evidence base is the limited number of children
with asthma in the cohort studies evaluated, both in recent studies and in the 2013 Pb ISA. This decreases
the statistical power to detect an association. Although less informative than the prospective cohort
studies due to a lack of temporality and a less relevant exposure window, recent cross-sectional NHANES
6-37
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analyses also reported null associations between childrens' BLLs and asthma (Wells et al.. 2014). eczema
(Wei et al.. 2019). and food allergies (Mener et al.. 2015) in much larger study populations. Results from
recent epidemiologic studies of allergen-specific and non-specific immunological biomarkers of
hypersensitivity in children and adults were less consistent than the generally null results for atopic
diseases, providing inconsistent evidence in both children and adults.
Recent toxicological evidence for effects of Pb exposure on biomarkers of allergic disease is
sparse and limited to two reports investigating cytokine levels in blood. Decreased IFN-y, a Thl cytokine
known to play a role in the resolution of asthma, was reported in a recent study. Pb exposure had no effect
on the levels of other cytokines that have been reported to play a role in allergic disease (i.e., GM-CSF,
IL-6, IL-10, and TNF-a). However, the value of these data for hazard identification is limited by two
factors. Changes in cytokine levels (particularly when measured in blood) can be associated with many
different types of tissues and toxicities and may reflect an immune response to tissue injury but not
necessarily an effect on or impairment of immune function. For this reason, cytokine secretion data (in the
absence of a stimulus) are considered supporting evidence for understanding mechanisms of immune
disruption, not as apical data. In addition, the utility of these data is further diminished by the lack of
additional studies corroborating these findings.
In summary, the collective body of evidence is suggestive of, but not sufficient to infer, a
causal relationship between Pb exposure and sensitization and allergic responses. Whereas a few
small prospective studies reviewed in the 2013 Pb ISA supported the presence of an association between
BLLs and incident asthma in children, recent prospective epidemiologic studies provide little evidence of
an association between exposure to Pb and atopic disease in children and inconsistent evidence for
immunological biomarkers involved in sensitization and allergic response. The recent epidemiologic
studies add considerable uncertainty to the line of evidence that previously provided support for the
'likely to be causal' determination in the 2013 Pb ISA. Differences in study designs and exposure
concentrations do not appear to explain the inconsistency in results of the more recent studies compared
to studies reviewed in the 2013 Pb ISA. While the epidemiologic evidence base for sensitization and
allergic response is inconsistent, there is consistent toxicological evidence that exposure to Pb increased
lymph node cell proliferation, increased production of Th2 cytokines such as IL-4, increased total serum
IgE antibody levels in serum, and misregulated inflammation in studies reporting BLL relevant to this
ISA. Biological plausibility for the associations observed in some epidemiologic studies is provided by
toxicological evidence that Pb exposure (1) promotes the production of Th2 cells and cytokines including
IL-4 and (2) increases total serum IgE levels in studies utilizing non-relevant routes of administration
(i.e., injection) and in studies either reporting high BLLs or those not reporting BLLs at all.
6-38
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Table 6-2 Summary of evidence that is suggestive of, but not sufficient to infer, a causal relationship
between Pb exposure and sensitization and allergic responses
Rationale for Causality K F./iH~n_„b Kpv Rpforpnrpcb Pb Biomarker Levels Associated with
Determination3 *ey tviaence *ey Terences Effects0
Consistent evidence from other Increased IL-4 production, decreased IFN- Fernandez-Cabezudo et al. (2007)
toxicological studies with y production in mice administered Pb in
relevant exposures drinking water for 16 wk
investigating immune functional
endpoints Increased IL-4 production in mice exposed lavicoli et al. (2006)
prenatally and postnatally
Increased total serum IgE antibody in mice Snyder et al. (2000)
exposed prenatally and postnatally to
0.1 mM Pb acetate for 2 wk
Mean BLL: 5 or 10 mM with BLL of 20.5
and 106.2 |jg/dL, respectively
0.02, 0.06, 0.11, 0.2, 40.00, and 400.0 ppm
with mean BLL of 0.83, 1.23, 1.59, 1.97,
11.86, and 61.48 |jg/dL, respectively
Mean BLL: 25.3 pg/dL
Inconsistent epidemiologic A limited number of studies reported
evidence for atopic disease positive but imprecise associations
provides limited coherence with between BLLs and asthma incidence and
toxicological evidence prevalence in children. Studies limited by
small number of cases
A limited number of recent studies with
lower BLLs reported null associations
between BLLs and asthma incidence and
prevalence in children
Generally null associations observed in
studies of other atopic diseases in
children, including eczema and food
allergies
Joseph et al. (2005)
Puqh Smith and Nriaqu (2011)
Pesce et al. (2021)
Wells et al. (2014)
See Section 6.4.2
Associations observed in stratified analysis
for participants with BLLs >5 and >10 pg/dL
Mean cord BLL: 1.45 pg/dL
Geometric Mean BLL: 1.13 pg/dL
Mean/Median BLL across studies:
1.01-1.75 pg/dL
Biological Plausibility Evidence that Pb (1) promotes T cell See Section 6.6
skewing leading to the production of Th2
cells and cytokines including IL-4, (2)
increased IgE levels, and (3) increased
inflammation in animals
BLL = blood lead level; IFN-y = interferon-gamma; IgE = immunoglobulin E; IL-4 = interleukin 4; Pb = lead.
"Based on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
'Describes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or inconsistencies.
References to earlier sections indicate where the full body of evidence is described.
Describes the Pb biomarker levels at which the evidence is substantiated.
6-39
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6.7.3
Causality Determination for Autoimmunity and Autoimmune Disease
In the 2013 Pb ISA, it was concluded "the evidence is inadequate to determine if there is a
causal relationship between Pb exposure and autoimmunity/' (U.S. EPA. 2013). This causality
determination was reached based on evaluation of a limited body of evidence that does not sufficiently
inform Pb-induced generation of autoantibodies with relevant Pb exposures. While elevated levels of
autoantibodies were reported in a single study of Pb-exposed battery workers with BLLs (10-40 (ig/dL)
(El-Fawal et al.. 1999), the internal validity and relevance of this study to this ISA is uncertain because of
a lack of adjustment for important confounders. In the only toxicology study available for the 2013 Pb
ISA with BLLs relevant to humans, autoantibodies were detected in rats following dietary administration
of Pb resulting in BLLs of 11-50 (ig/dL (El-Fawal et al., 1999).
Recent epidemiologic studies of autoimmunity are limited in number and examine disparate
outcomes (Joo et al., 2019; Kamycheva et al„ 2017). Neither study observed evidence supporting an
association between Pb exposure and autoimmunity. Although Kamycheva et al. (2017) reported an
inverse association between BLLs and seropositivity for Celiac Disease, the cross-sectional study design
does not preclude reverse causality, whereby the association may result from reduced absorption of Pb
rather than a protective effect of Pb exposure. Only one recent toxicology study was available for this
assessment. In that study, Fang et al. (2012) reported that administration of Pb acetate in drinking water
for 42 days (BLL = 18.48 (ig/dL) had no effect on the suppressive properties of Tregs isolated from adult
male Sprague Dawley rats.
In summary, the collective body of evidence remains inadequate to infer the presence or
absence of a causal relationship between Pb exposure and autoimmunity and autoimmune disease.
This determination is based on the limited number of epidemiologic and toxicological studies and the
disparate outcomes examined therein, which make it difficult to draw conclusions about the nature of the
relationship. The evidence available to date does not indicate a relationship between exposure to Pb and
autoimmunity and autoimmune disease.
6-40
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Table 6-3 Summary of evidence that is inadequate to determine the presence or absence of a causal
relationship between Pb exposure and autoimmunity and autoimmune disease
Rationale for Causality K Fviri_nr_b Kpu Rpfprpnrp«.b Pb Biomarker Levels Associated with
Determination3 *ey tviaence *ey inferences Effects*
Limited toxicological evidence A study in rats shows generation of El-Fawal et al. (1999) BLL: 11-50 |jg/dL
for increased autoantibodies autoantibodies with relevant adult-only
oral Pb exposure for 4 d. Several other
studies have Pb exposure concentrations
and/or exposure routes
(e.g., intraperitoneal) with uncertain
relevance to humans
Coherence from a limited
number of epidemiologic
studies for increased
autoantibodies at high BLLs
Evidence for increased autoantibodies in El-Fawal et al. (1999)
Pb-exposed workers with high BLL and
limited consideration for potential
confounding, including other workplace
exposures
BLL: 10-40 pg/dL
Lack of coherence from
epidemiologic studies of
autoimmune disease
Limited number of epidemiologic studies
reported null or associations between
BLLs and autoimmune disease
Kamvcheva et al. (2017)
Joo et al. (2019)
Limited evidence for biological Administration of Pb for 42 d had no
plausibility effect on Treg activity in rats
Fang et al. (2012)
BLL: 18.48 pg/dL
BLL = blood lead level; d = day; Pb = lead; Treg = regulatory T cells.
"Based on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
'Describes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
Describes the Pb biomarker levels at which the evidence is substantiated.
6-41
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6.8
Evidence Inventories - Data Tables to Summarize Study Details
Table 6-4 Epidemiologic studies of exposure to Pb and immunosuppression
Referent^ and Study study Popu|atjon
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Host Resistance
tEqqers et al. (2018) NHANES
n: 18626
United States
2001-2004
Cross-Sectional
General population;
>1 yr old
Blood
Blood Pb was measured in
venous whole blood using
GFAAS (2001-2002) and ICP-
MS (2003-2004)
Age at measurement:
>1 yr old
Median: 1.4 |jg/dL
75th: 2.3 pg/dL
Maximum: 68.9 pg/dL
Q1
Q2
Q3
Q4
<0.91 pg/dL
0.91-1.4 pg/dL
1.41-2.3 pg/dL
>2.3 pg/dL
Prevalence of MRSA and
MSSA colonization
Colonization by S. aureus
tested using nasal swabs
and standard culture-
based procedures
Age at Outcome:
>1 yr old
Age, sex, race, income,
smoking, iron, calcium, and
Vitamin C
ORs
MRSA Colonization:
Q1: Reference
Q2
Q3
Q4
1.52 (0.83, 2.76)
1.56 (0.75, 3.24)
1.82 (0.81, 4.1)
MRSA Colonization:
Q1
Q2
Q3
Q4
Reference
1.07 (0.95, 1.21)
1.1 (0.94, 1.28)
0.91 (0.76, 1.09)
6-42
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tKrueqer and Wade NHANES
(2016)
United States
1999-2012
Cross-Sectional
n: 18,425 (7. gondii)
17,389 (hepatitis B),
5,994 (H. pylori)
General population; >3
yr old (H. pylori), >6 yr
old (T. gondii and
HBV)
Blood
Blood Pb was measured in
venous whole blood using ICP-
MS
Age at measurement:
>3 yr old (H. pylori), >6 yr old
(T. gondii and HBV)
Geometric mean: 1.5 pg/dL
Seropositivity for T. gondii,
H. pylori, and hepatitis B
Serum tested for T. gondii
and H. pylori IgG
antibodies using an ELISA
and HBc ELISA was used
to detect total antibodies
against hepatitis B core
antigen
Age at Outcome:
>3 yr old (H. pylori), >6 yr
old (T. gondii and HBV)
Age, sex, race/ethnicity,
country of birth, family
income, self-reported health,
tap water source, household
crowding, NHANES cycle,
and use of illicit intravenous
drugs
ORs
H. pylori Seropositivity:
I.09 (1.05, 1.13)
T. gondii Seropositivity:
1.10 (1.06, 1.14)
Hepatitis B
Seropositivity:
1.08 (1.03, 1.13)
tFeiler et al. (2020)
Rochester, NY
United States
2012-2017
Case-control
n: 2,663 (full sample);
617 (reduced sample)
Test-negative case-
control study of
children <4 yr old
tested for
influenza/RSV
Blood
Blood Pb measured in venous or
capillary whole blood samples
using GFAAS. When multiple
measurements were available
Age at measurement:
Between 6 mo and 4 yr
Mean: NR
-60% of children had peak BLLs
<1 |jg/dL; 5% had peak BLLs
>5 |jg/dL
Influenza and RSV
diagnosis
Nasopharyngeal swab
samples tested for
influenza or RSV by PCR
Age at Outcome:
<4 yr old
Full sample: age, sex, race,
ethnicity, insurance status,
and respiratory season.
Reduced sample: Same as
full, plus maternal age,
parity, feeding type, maternal
smoking, and area-level
poverty, unemployment,
education, and housing built
before 1980
ORs
Influenza
<1 pg/dL: Reference
1-3: 1.52 (0.69, 3.37)
>3: 1.12 (0.45, 2.82)
RSV
<1 pg/dL: Reference
1-3: 0.97 (0.56, 1.66)
>3: 0.9 (0.5, 1.62)
6-43
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RefereDCesignnd Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tParketal. (2020)
n: 2625
Blood
H. pylori infection
Age, smoking, drinking, BMI,
and diabetes, exercise
ORs
Hwasun
South Korea
Patients >20 yr old
undergoing
Blood Pb measured in whole
blood using GFAAS
H. pylori infection
confirmed histologic
H. pylori Infection
2014-2016
gastrointestinal
Age at measurement:
examination using
endoscopy
>20 yr old
Giemsa staining of
Men: 1.05 (1.03, 1.08)
Women: 1.06 (1.00, 1.13)
Cross-sectional
Mean:
Men: 3.15 |jg/dL; Women:
abnormal lesions
identified during
endoscopy
2.19 |jg/dL
Age at Outcome:
>20 yr old
Vaccine Antibody Response
tDi Lenardo et al.
Venda Health
Blood
Measles, Tetanus, and H.
Maternal age, HIV status,
ORs for odds of being
(2020)
Examination of
influenzae type B IgG
duration of breast feeding
below protective cut
Mothers, Babies and
Blood Pb measured in triplicate
titers
point
Limpopo
their Environment
in whole blood using ICP-MS
South Africa
2012-2013
n: 425
Age at measurement:
1 yr
Serum IgG specific to
measles, tetanus, and Hib
Measles IgG Levels:
Women recruited
measured by ELISA
1.00 (0.77, 1.31)
Cohort
when presenting for
Median: 1.9 |jg/dL
delivery. Children were
75th: 2.8 pg/dL
Age at Outcome:
Tetanus IgG Levels:
excluded if they did not
3.5 yr
receive measles,
1.13 (1.02, 1.26)
tetanus, and Hib
immunizations
Hib IgG Levels:
0.99 (0.89, 1.11)
6-44
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tJusko et al. (2019) NHANES
United States
1999-2004
Cross-Sectional
n: 7005
General population;
children 6-17 yr old.
Percent unvaccinated
not reported. MMR
vaccine schedule
between 1999 and
2004 was:
1st dose: 12-18 mo;
2nd dose: 4-6 yr; and
Catch-up 2nd dose by
11-12 yr
Blood
Blood Pb was measured in
venous whole blood using ICP-
MS
Age at measurement:
6-17 yr old
Mean: 1.4 pg/dL
Median: 1.0 pg/dL
Measles, Mumps, and
Rubella Antibody Levels
Measles and Rubella
antigen-specific IgG
levels were determined
using an ELISA; Mumps
antigen-specific IgG
levels were determined
via Wampole Mumps IgG
test
Age at Outcome:
6-17 yr old
Sex, age, race/ethnicity,
family poverty-income ratio,
and NHANES cycle
% Change
Anti-Measles IgG Levels:
-2.75 (-5.10, -0.41)
Anti-Mumps IgG Levels:
-2.07 (-3.87, -0.24)
Anti-Rubella IgG Levels:
0.00 (-2.58, 2.65)
tWelch et al. (2020) n: 502
Munshiganj and
Pabna
Bangladesh
2008-2011 enrollment
(follow-up through 5 yr
of age)
Cohort
Pregnant women with
singleton pregnancies
recruited and children
followed through 5 yr
of age
Blood
Cord blood Pb measured using
ICP-MS; Blood Pb measure in
capillary samples using portable
Lead-Care II instruments
Age at measurement:
At birth, 20-40 mo and 4-5 yr
Median:
Pregnancy: 3.1 |jg/dL;
Toddler: 6.4 |jg/dL;
Early Childhood: 4.7 |jg/dL
Serum vaccine antibody
concentrations (diphtheria
and tetanus)
Serum diphtheria and
tetanus antibodies
measured using an ELISA
Age at Outcome:
5 yr old
Maternal education,
breastfeeding duration, and
child sex
% Change in Median
Antibody Concentration
Cord BLLs
Diphtheria:
0.97 (-1.11, 3.05)
Tetanus:
1.54 (-0.17, 3.24)
BLLs
75th:
Pregnancy: 5.6 |jg/dL;
Toddler: 10.0 pg/dL;
Early Childhood: 7.0 pg/dL
Diphtheria:
-0.96 (-3.26, 1.33)
Tetanus:
0.33 (-2.36, 3.02)
6-45
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tXu etal. (2015)
Shantou
China
2011-2013
Cross-sectional
n: 490
Hepatitis B vaccinated
children 3-7 yr old
from two kindergartens
(one near an e-waste
facility, and the other
in a matched reference
area)
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
3-7 yr old
Geometric Mean:
Reference kindergarten:
6.05 |jg/dL; Exposed (e-waste)
kindergarten: 6.76 |jg/dL
Hepatitis B surface
antibody levels
Blood plasma HBsAb titer
was measured by ELISA
Age at Outcome:
3-7 yr old
Age and sex (areas matched
on traffic density, population,
SES, lifestyle, and cultural
background)
Change in HBsAb titers
(S/CO)
2011 Sample:
-0.45 (-0.49, -0.40)
2012 Sample:
-0.37 (-0.40, -0.33)
WBCs and Cytokines
tCao etal. (2018)
Guiyu and Haojiang
China
2014
Cross-Sectional
n: 118
Children 3-7 yr old at
two kindergartens (one
near an e-waste
facility, and the other
in a matched reference
area)
Blood
Pb measured in venous whole
blood using GFAAS
Age at measurement:
3-7 yr
Median: Reference kindergarten:
3.6 |jg/dL
Exposed (e-waste) kindergarten:
5.1 |jg/dL
T cell subpopulations, IL-
2, IL-7, IL-15 levels
T cell subpopulations
measured in whole blood
using flow cytometry;
Serum cytokines
measured using the
ProcartaPlex Human
Cytokine Chemokine
Panel 1A
Age at Outcome:
3-7 yr
Age and sex (areas matched
on traffic density, population,
SES, lifestyle, and cultural
background)
Change in percentage
of T cells
CD4+ Tn
-0.59 (-1.07, -0.12)
CD4+ Tcm
0.49 (0.10, 0.88)
6-46
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tChen etal. (2021) n: 486
Shantou
China
Nov.-Dec. 2018
Cross-sectional
Pre-school children
(aged 2-6) from two
towns with similar
but different Pb
exposure
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
2-6 yr
Median:
Exposed: 4.51 |jg/dL;
Reference: 3.98 |jg/dL
75th:
Exposed: 5.67 |jg/dL,
Reference: 4.84 |jg/dL
WBC, neutrophil, and
monocyte counts
WBCs, neutrophils, and
monocytes measured in
venous whole blood
Age at Outcome:
2-6 yr
Gender, age, BMI, e-waste
contamination w/ in 50 m of
residence, residence as
workplace, distance of
residence from road, family
member daily smoking,
monthly household income,
maternal work associated w/
e-waste, duration of outdoor
play, child contact w/ e-
waste, washing hands
before eating, nail biting
habit, chewing pencil habit,
yearly canned food
consumption, yearly
fruit/vegetable consumption,
yearly iron rich food
consumption, yearly marine
product consumption, and
yearly salted food
consumption
ln(WBC count)
0.006 (0.001, 0.012)
ln(Monocyte count)
0.006 (-0.001, 0.013)
ln(Neutrophil count)
0.009 (0, 0.018)
tDai etal. (2017)
Shantou
China
Cross-sectional
n: 484
Children 2-6 yr old
randomly sampled
from volunteers at two
kindergartens (one
near an e-waste
facility, and the other
in a matched reference
area)
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
2-6 yr old
Q1
Q2
Q3
Q4
<3.78 |jg/dL
3.78-5.22 |jg/dL
5.23-7.00 |jg/dL
>7.00 |jg/dL
Erythrocyte CR1
expression measured
using flow cytometry
Age at Outcome:
2-6 yr old
Age, gender, paternal and
maternal education level,
and family income
Mean Difference in
Erythrocyte CR1
Expression
Q1
Q2
Q3
Q4
Reference
-0.07 (-0.23, 0.08)
-0.04 (-0.20, 0.11)
-0.16 (-0.32, -0.01)
6-47
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tHuo etal. (2019)
Shantou
China
NR
Cross-sectional
n: 267
Children 2-7 yr old at
two kindergartens (one
near an e-waste
facility, and the other
in a matched reference
area)
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
2-7 yr old
Median:
Reference kindergarten:
4.4 |jg/dL; Exposed (e-waste)
kindergarten: 6.5 |jg/dL
75th:
Reference kindergarten:
5.6 |jg/dL; Exposed (e-waste)
kindergarten: 8.2 |jg/dL
IFN-y, IL-113, and IL-
12p70
Serum cytokine measured
using the ProcartaPlex
Human Cytokine
Chemokine Panel 1A
Age at Outcome:
2-7 yr old
Age and sex (areas matched
on traffic density, population,
SES, lifestyle, and cultural
background)
Per natural log increase in
erythrocyte Pb
IL-ip pg/ml
0.08 (-0.01, 0.17)
IL-12p70 pg/ml
0.99 (0.53, 1.44)
ifn-y pg/mi
1.43 (0.57, 2.30)
tLi etal. (2018)
Hubei and Hunan
Provinces
China
2012-2017
Cross-Sectional
Blood Lead
Intervention Program
n: 758
Children Ages 5-8 yr
recruited from 4
counties in 2
provinces. One county
in each province had
high environmental Pb
levels (battery plant
and mining)
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
5-8 yr old
Geometric mean: 8.24 |jg/dL
75th: 13.51 pg/dL
90th: 18.77 pg/dL
95th: 21.82 pg/dL
WBC count
Hematological
parameters were
analyzed by an
automated hematology
analyzer (BC-5800;
Mindray, Shenzhen,
China) with quality control
processes.
Age at Outcome:
5-8 yr old
Age, gender, BMI,
environmental lead exposure
level, and serum iron, zinc,
and calcium
OR
Decreased WBC count
(<4 x 109/L)
1 (0.905, 1.105)
6-48
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RefereDCesignnd Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tWerder et al. (2020) Gulf Long-Term
Follow-up Study
n: 214
Pb measure in blood using solid- IL-6, IL-8, IL-113, TNF-a Age, race, alcohol
Gulf Region
United States
2012-2013
Cross-sectional
Non-smoking >30 yr
old male oil spill
response workers and
oil spill safety trainees
with no history of liver
disease or heavy
alcohol use
phase micro-extraction with gas
chromatography/mass
spectrometry
Age at measurement:
>30
Mean: 1.82 |jg/dL
Cytokeratin 18 (CK18
M65 and CK18 M30)
Age at Outcome:
>30
consumption, serum
cotinine, BMI, diabetes
diagnosis, and education
pg/mL change (obese
participants)
IL-6
169.6 (119.8, 219.4)
IL-8
360.9 (246.2, 475.6)
IL-1 p
76.3 (63.6, 89.0)
TNF-p
1.1 (-1.5, 3.6)
6-49
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tZhanq et al. (2020) n: 147
Shantou
China
Cross-sectional
Children 3-7 yr old at
two kindergartens (one
near an e-waste
facility, and the other
in a matched reference
area)
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
3-7 yr old
Median:
Reference kindergarten:
2.3 |jg/dL; Exposed (e-waste)
kindergarten: 3.7 |jg/dL
Neutrophils, monocytes,
lymphocytes, IL-113, IL-6,
IL-8, IL-10, and TNF-a
Immune cells measured
in whole blood using an
automated blood cell
analyzer; Serum
cytokines measured using
the ProcartaPlex Human
Cytokine Chemokine
Panel 1A
Age at Outcome:
3-7 yr old
Gender, age, BMI, e-waste
contamination w/ in 50 m of
residence, residence as
workplace, distance of
residence from road, family
member daily smoking,
maternal work associated w/
e-waste, child contact w/ e-
waste, washing hands
before eating, milk
consumption frequency, and
ventilation of house
Per natural log increase in
erythrocyte Pb
In(Neutrophils)
0.20 (0.00, 0.39)
In(Monocytes)
0.02 (-0.14, 0.18)
In(Lymphocytes)
-0.05 (-0.24, 0.16)
ln(IL-1b)
0.19 (-0.08, 0.45)
ln(IL-6)
0.33 (0.04, 0.62)
ln(IL-8)
0.05 (-0.28, 0.37)
ln(IL-10)
0.08 (-0.29, 0.44)
In(TNF-a)
-0.18 (-0.44, 0.08)
BLL = blood lead level; BMI = body mass index; CD = cluster of differentiation; CI = confidence interval; CK = cytokeratin; CR1 = complement receptor type 1; e-waste = electronic-waste;
ELISA = enzyme-linked immunosorbent assay; GFAAS = graphite furnace atomic absorption spectrometry; HBc = Hepatitis B core; HBsAb = Hepatitis B surface antigen; HBV = Hepatitis B
virus; Hib = Haemophilus influenzae type B; ICP-MS = inductively coupled plasma mass spectrometry; Ig- = immunoglobulin type; IL = interleukin type; IFN-g = interferon-gamma;
In = natural log; mo = month(s); MRSA = methicillin-resistant Staphylococcus aureus; NHANES = National Health and Nutrition Examination Survey; NR = not reported; OR = odds ratio;
Pb = lead; PCR = polymerase chain reaction; RSV = respiratory syncytial virus; S/CO = signal to cut-off; SES = socioeconomic status; SPT = skin prick test; TNF-a = tumor necrosis factor
alpha; WBC = white blood cell; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb level or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
fStudies published since the 2013 Pb ISA.
6-50
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Table 6-5
Animal toxicological studies of delayed-type hypersensitivity responses
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported Endpoints
(pg/dL)a Examined
Fanq et al. (2012)
Rat (Sprague Dawley)
Control (vehicle), M,
n = 20
300 ppm Pb, M, n = 20
23-25 d to 65-
67 d
Dosing solutions were changed twice
per wk
4.48 |jg/dL for 0 ppm DTH
18.48 |jg/dL for
300 ppm - d 65-67
BLL = blood lead level; d = day; DTH = delayed-type hypersensitivity; M = male; MMR = measles, mumps, and rubella; Pb = lead; ppm = parts per million; wk = week
alf applicable, reported values for BLL were converted to
-------
Table 6-6 Animal toxicological studies of antibody response
Study Species (Stock/Strain), n, Timing of Exposure Exposure Details BLL as Reported Endpoints
Sex a H (Concentration, Duration) (HSJ'dL) Examined
Yathapu et al.
(2020)
Rat (Sprague Dawley)
Control (vehicle)
M/F, n = 32 (16/16)
PND 54- PND 82
Weanling rats (PND 21) were acclimated
to the facility for 5 days before being
divided into two groups (n = 16) to begin
a 28-day long Fe deficiency diet. After
28 days, the rats were exposed to Pb or
control diet (n = 16). At this point
(PND 82), blood was collected from rats
before immunization with TT (n = 8)
followed by two boosters administered in
2-wk intervals. Vaccine response was
evaluated 2 wk later
2.1 ± 1.0 pg/dL for
0 mg/4 mL/kg,
16.1 ± 5.5 pg/dL for
25 mg/4 mL/kg - PND 82,
Control diet
1.9 ± 0.7 pg/dL for
0 mg/4 mL/kg
41.6 ± 10.2 pg/dL for
25 mg/4 mL/kg - PND 82,
Iron deficiency diet
Vaccine
response,
Antigen-specific
antibodies
BLL = blood lead level; Fe = iron; M/F = male/female; Pb = lead; PND = postnatal day; TT = tetanus toxoid.
alf applicable, reported values for BLL were converted to mg/dL using WebPlot Digitizer (https://apps.automeris.io/wpd/') and are shown in parenthesis.
Table 6-7 Animal toxicological studies of ex vivo white blood cell function
Study
Species (Stock/Strain), n,
Sex
Timing of Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported
(Mg/dLf
Endpoints
Examined
Fang et al. (2012)
Rat (Sprague Dawley)
Control (vehicle), M, n = 20
300 ppm Pb, M, n = 20
23-25 d to 65-67 d
Dosing solutions were changed twice
per wk.
4.48 pg/dL for 0 ppm, Tregcell
18.48 pg/dL for suppression
300 ppm — d 65-67 assay
6-52
-------
Yathapu et al. (2020)
Rat (Sprague Dawley)
Control (vehicle), M/F, n = 32
(16/16)
500 ppm Pb, M/F, M/F,
n = 32 (16/16)
PND 54 - PND 82
Weanling rats (PND 21) were
acclimated to the facility for 5 days
before being divided into two groups
(n = 16) to begin a 28-day long Fe
deficiency diet. After 28 days, the rats
were exposed to Pb or control diet
(n = 16). At this point (PND 82), blood
was collected from rats before
immunization with TT (n = 8) followed
by two boosters administered in 2-wk
intervals. Vaccine response was
evaluated 2 wk later.
2.1 ± 1.0 pg/dL for
0 mg/4 mL/kg
16.1 ± 5.5 pg/dL for
25 mg/4 mL/kg -
PND 82, Control diet
1.9 ± 0.7 pg/dL for
0 mg/4 mL/kg
41.6 ± 10.2 pg/dL for
25 mg/4 mL/kg -
PND 82, Iron
deficiency diet
Spleen cell
proliferation
BLL = blood lead level; d = day; Fe = iron; M/F = male/female; Pb = lead; PND = postnatal day; ppm = parts per million; Treg = regulatory T cells; TT = tetanus toxoid; wk = week.
alf applicable, reported values for BLL were converted to mg/dL using WebPlot Digitizer (https://apps.automeris.io/wpd/) and are shown in parenthesis.
6-53
-------
Table 6-8
Animal toxicological studies of immune organ pathology
Study
Species (Stock/Strain), n, Timing of Exposure Details
Sex Exposure (Concentration, Duration)
BLL as Reported
(ng/dL)a
Endpoints
Examined
Corsetti etal. (2017) Mouse (C57BJ)
Control (vehicle), M, n
200 ppm Pb, M, n = 8
30-75 d
Mice were exposed via drinking water
for 45 consecutive days. Control
animals were exposed to drinking
water containing acetic acid (1 mL/L)
<5 |jg/dL for 0 ppm
21.6 |jg/dLfor200 ppm
Spleen
histopathology
Dumkova et al. (2017) Mouse (ICR) NR
Control (vehicle), F, n = 10
1.23 x 10s particles/cm3 Pb,
F, n = 10
Mice were exposed continuously
(24 hr/d, 7 d/wk) for 6 wk. Control
animals were exposed to the same air
as the treated group without the
addition of Pb nanoparticles. The
investigators pooled animals from two
independent experiments, each with
five animals per treatment
11 ng/g forO * 10s
particles/cm3 Pb
(1.166 |jg/dL)
132 ng/g for 1.23 x 10®
particles/cm3 Pb
(13.992 |jg/dL)
Spleen
histopathology
Dumkova et al.
(2020b)
Mouse CD-1 (ICR) NR
Control (vehicle), F, n = 10
(2 wk, 6 wk, 11 wk)
2.23 x 10® NPs/cm3 PbO
NP, F, n = 10 (2 wk, 6 wk,
11 wk)
2.23 x 10s NPs/cm3 PbO NP
recovery, F, n = 10 (6 wk
PbO NP, 5 wk clean air)
Mice (unknown age) were exposed to <3 ng/g for 0 PbO
clean air or PbO NPs 24 hr/d 7 d/wk
for 2 wk, 6 wk, or 11 wk. a recovery
group was exposed to PbO NPs for
6 wk and then clean air for 5 wk
(11 wk total)
NPs/cm3(<0.3 |jg/dL)
104 ng/g for 2.23 * 10®
N Ps/cm3 - 2 wk
(10.4 |jg/dL)
<3 ng/g for 0 PbO
N Ps/cm3 - 6 wk
(<0.3 |jg/dL)
148 ng/g for 2.23 x 10®
N Ps/cm3 - 6 wk
(14.8 |jg/dL)
Spleen
histopathology
<3 ng/g for 0 PbO
N Ps/cm3 -11 wk
(<0.3 |jg/dL)
6-54
-------
Species (Stock/Strain), n, Timing of Exposure Details BLL as Reported Endpoints
Sex Exposure (Concentration, Duration) (ng/dL)a Examined
Dumkova et al. MouseCD-1 (ICR)
(2020a) Control (vehicle), F, n = 10
68.6 |jg/m3 Pb, F, n = 10
174 ng/g for 2.23 * 10®
N Ps/cm3 -11 wk
(17.4 |jg/dL)
<3 ng/g for 0 PbO
NPs/cm3 (<0.3 |jg/dL)
27 ng/g - recovery (6 wk
PbO NP, 5 wk clean air)
(2.7 pg/dL)
6-8 wk old mice
exposed for 3 d,
2 wk, 6 wk, or
11 wk
40 ng/g for 68.6 pg/m3 Pb
- 2 wk (4.0 pg/dL)
47 ng/g for 68.6 pg/m3 Pb
- 6 wk (4.7 pg/dL)
85 ng/g for 68.6 pg/m3 Pb
-11 wk (8.5 pg/dL)
10 ng/g for 68.6 pg/m3 Pb
- 6 wk exposure plus
5 wk clean air
(1.0 pg/dL)
Mice were exposed to Pb for 3 d,
2 wk, 6 wk, or 11 wk. To assess
recovery, a separate group of mice
were exposed for 11 wk followed by
5 wk of clean air. Control group was
exposed to filtered air
<0.3 ng/g for control at all Spleen
timepoints (d 3, 2 wk, histopathology
6wk, 11 wk) (<0.3 pg/dL)
31 ng/g for 68.6 pg/m3 Pb
-d 3 (3.1 pg/dL)
6-55
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported
(ng/dL)a
Endpoints
Examined
Smutna et al. (2022)
Mouse CD-1 (ICR)
Control (vehicle), F, n = 10
0.956 pg/m3 Pb, F, n = 10
6-8 wk old mice
exposed for 11 wk
Mice were exposed to Pb for
11 wk. Control group was exposed to
filtered air
<0.003 ± 0.001 ng/g for
control at 11 wk
(0.318 ±0.106 pg/dL)
0.171 ± 0.012 ng/g for
0.956 pg/m3 Pb -11 wk
(18.126 ± 1.272 pg/dL)
Spleen
histopathology
BLL = blood lead level; d = day; F = female; Pb = lead; PbO nanoparticles = lead oxide nanoparticles; ppm = parts per million; wk = week.
alf applicable, reported values for BLL were converted to mg/dL using WebPlot Digitizer (https://apps.automeris.io/wpd/) and are shown in parenthesis.
Table 6-9 Animal toxicological studies of immunoglobulin levels
Study Species (Stock/Strain), n, Timing of Exposure Details BLL as Reported (ug/dL)a fndpoints
y Sex Exposure (Concentration, Duration) p ' Examined
Yathapu et al.
(2020)
Rat (Sprague Dawley)
Control (vehicle)
M/F, n = 32 (16/16)
500 ppm Pb, M/F, n = 32
(16/16)
PND 54-PND 82
Weanling rats (PND 21) were
acclimated to the facility for 5 days
before being divided into two groups
(n = 16) to begin a 28-day long Fe
deficiency diet. After 28 days, the rats
were exposed to Pb or control diet
(n = 16). At this point (PND 82), blood
was collected from rats before
immunization with TT (n = 8) followed
by two boosters administered in 2- wk
intervals. Vaccine response was
evaluated 2 wk later.
2.1 ± 1.0 pg/dL for
0 mg/4 mL/kg - PND 82,
Control Diet
16.1 ± 5.5 pg/dL for
25 mg/4 mL/kg - PND 82,
Control diet
1.9 ± 0.7 pg/dL for
0 mg/4 mL/kg - PND 82, Iron
deficiency diet
Immunoglobulin
levels
41.6 ± 10.2 pg/dL for
25 mg/4 mL/kg - PND 82,
Iron deficiency diet
BLL = blood lead level; Fe = iron; M/F = male/female; Pb = lead; PND = postnatal day; TT = tetanus toxoid.
alf applicable, reported values for BLL were converted to mg/dL using WebPlot Digitizer (https://apps.automeris.io/wpd/) and are shown in parenthesis.
6-56
-------
Table 6-10
Animal toxicological studies of immune organ weight
Study
Species (Stock/Strain), n, Timing of
Sex Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (|jg/dL)a
Endpoints
Examined
Amos-Kroohs et al.
(2016)
Rat (Sprague Dawley)
Control (vehicle), M/F, n = 4
(2/2)
1 mg/kg Pb, M/F, n = 16
(8/8)
PND4-PND28
10 mg/kg Pb,
(8/8)
M/F, n = 16
Male and female rats were gavaged
every other day from PND 4 to
PND 10, 18, or 28. Starting on
PND 4, ISO offspring were isolated
from their dam individually for 4 hr.
Control animals remained with their
dam throughout this period. On
PND 11, 19, or 29, subsets within
each group were subjected to acute
stressor (shallow water stressor for 0,
30, or 60 min) or left
undisturbed. Control animals were
gavaged with vehicle containing
anhydrous sodium acetate (0.01 M)
1.19 |jg/dL for 0 mg/kg
2.73 |jg/dL for 1 mg/kg
9.15 |jg/dL for 10 mg/kg ¦
PND 29 w/o ISO stress
1.31 pg/dL for 0 mg/kg,
4.55 pg/dL for 1 mg/kg
17.1 pg/dL for 10 mg/kg ¦
PND 29 w/ ISO stress
Spleen weight,
Thymus
weight
Corsetti et al. (2017) Mouse (C57BJ)
Control (vehicle), M, n
200 ppm Pb, M, n = 8
d 30-d 75 Mice were exposed via drinking water <5 pg/dL for 0 ppm
for 45 consecutive days. Control
animals were exposed to drinking
water containing acetic acid (1 mL/L)
21.6 pg/dL for 200 ppm
Spleen weight
Dumkova et al. (2017) Mouse (ICR) NR
Control (vehicle), F, n = 10
1.23 x 10s particles/cm3 Pb,
F, n = 10
Mice were exposed continuously 11 ng/g forO * 10s Spleen weight
(24 h/d, 7 d/wk) for 6 wk. particles/cm3 Pb (1.166 pg/dL)
Control animals were exposed to the „„R
m . . i ... , 32 nq/q for 1 23 * 10s
same air as the treated group without a a 3
the addition of Pb nanoparticles.
The investigators pooled animals
from two independent experiments,
each with five animals per treatment
particles/cm3 Pb
(13.992 pg/dL)
6-57
-------
Study
Species (Stock/Strain), n, Timing of
Sex Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (|jg/dL)a
Endpoints
Examined
Dumkova et al. Mouse CD-1 (ICR) NR
(2020b) Control (vehicle), F, n = 10
(2 wk, 6 wk, 11 wk)
2.23 x 10® NPs/cm3 PbO
NP, F, n = 10 (2 wk, 6 wk,
11 wk)
2.23 x 10® NPs/cm3 PbO
NP recovery, F, n = 10
(6 wk PbO NP, 5 wk clean
air)
Mice (unknown age) were exposed to
clean air or PbO NPs 24 hr/d, 7 d/wk
for 2 wk, 6 wk, or 11 wk. a recovery
group was exposed to PbO NPs for
6 wk and then clean air for 5 wk
(11 wk total)
<3 ng/g for 0 PbO NPs/cm3-
2 wk (<0.3 |jg/dL)
104 ng/g for 2.23 x 10®
NPs/cm3 - 2 wk (10.4 |jg/dL)
<3 ng/g for 0 PbO NPs/cm3 -
6wk (<0.3 |jg/dL)
148 ng/g for 2.23 * 10®
NPs/cm3 - 6 wk (14.8 pg/dL)
<3 ng/g for 0 PbO NPs/cm3 -
11 wk (<0.3 pg/dL)
Spleen weight
174 ng/g for 2.23 x 10®
NPs/cm3 - 11 wk (17.4 pg/dL)
<3 ng/g for 0 PbO NPs/cm3
(<0.3 pg/dL)
27 ng/g - recovery (6 wk PbO
NP, 5 wk clean air) (2.7 pg/dL)
Dumkova et al.
Mouse CD-1 (ICR)
6-8 wk old mice
Mice were exposed to Pb for 3 d,
<0.3 ng/g for control at all Spleen weight
(2020a)
Control (vehicle), F, n = 10
exposed for 3 d,
2 wk, 6 wk, or 11 wk. To assess
timepoints (d 3, 2 wk, 6 wk,
2 wk, 6 wk, or
recovery, a separate group of mice
11 wk) (<0.3 pg/dL)
68.6 pg/m3 Pb, F, n = 10
11 wk
were exposed for 11 wk followed by
5 wk of clean air. Control group was
exposed to filtered air
31 ng/g for 68.6 pg/m3 Pb -
d 3
(3.1 pg/dL)
40 ng/g for 68.6 pg/m3 Pb -
2 wk (4.0 pg/dL)
6-58
-------
Study Species (Stock/Strain), n, Timing of Exposure Details BLL as Reported (ug/dL)a fndpoints
Sex Exposure (Concentration, Duration) Examined
47 ng/g for 68.6 |jg/m3 Pb -
6 wk (4.7 |jg/dL)
85 ng/g for 68.6 |jg/m3 Pb -
11 wk (8.5 |jg/dL)
10 ng/g for 68.6 |jg/m3 Pb -
6 wk exposure plus 5 wk clean
air (1.0 |jg/dL)
Smutna et al. (2022) Mouse CD-1 (ICR)
6-8 wk old mice
Mice were exposed to Pb for 11 wk.
<0.003 ± 0.001 ng/g for control
Spleen
Control (vehicle), F, n = 10
exposed for
Control group was exposed to filtered
at 11 wk (0.318 ±0.106 pg/dL)
histopathology
11 wk
air
0.956 |jg/m3 Pb, F, n = 10
0.171 ± 0.012 ng/g for
0.956 pg/m3 Pb - 11 wk
(18.126 ± 1.272 pg/dL)
Graham et al. (2011) Rat (Sprague Dawley), PND 4-PND 28 Dosed every other day. Control 0.267 |jg/dL for 0 mg/kg, Spleen weight,
animals were gavaged with vehicle 3.27 |jg/dL for 1 mg/kg, Thymus
containing anhydrous sodium acetate 12.5 |jg/dL for 10 mg/kg - weight
(0.01 M) PND 29
6-59
-------
Study Species (Stock/Strain), n, Timing of Exposure Details BLL as Reported (ug/dL)a fndpoints
Sex Exposure (Concentration, Duration) Examined
Groups:
PND 11
Control (vehicle), M/F,
n = 192 (96/96)
1 mg/kg Pb, M/F, n = 192
(96/96)
10 mg/kg Pb, M/F, n = 191
(96/95)
PND 19
Control (vehicle), M/F,
n = 191 (96/95)
1 mg/kg Pb, M/F, n = 191
(96/95)
10 mg/kg Pb, M/F, n = 192
(96/96)
PND 29
Control (vehicle), M/F,
n = 192 (96/96)
1 mg/kg Pb, M/F, n = 192
(96/96)
10 mg/kg Pb, M/F, n = 192
(96/96)
Graham et al. (2011) Rat (Spraque Dawlev)
PND 4-PND 28 Dosed every other day. Control
0.267 |jg/dL for 0 mg/kg -
Spleen weight,
animals were gavaged with vehicle
PND 29
Thymus
containing anhydrous sodium acetate
weight
(0.01 M)
3.27 |jg/dL for 1 mg/kg -
PND 29
12.5 |jg/dL for 10 mg/kg -
PND 29
6-60
-------
Study Species (Stock/Strain), n, Timing of Exposure Details BLL as Reported (ug/dL)a fndpoints
Sex Exposure (Concentration, Duration) Examined
Groups:
PND 11
Control (vehicle), M/F,
n = 192 (96/96)
1 mg/kg Pb, M/F, n = 192
(96/96)
10 mg/kg Pb, M/F, n = 191
(96/95)
PND 19
Control (vehicle), M/F,
n = 191 (96/95)
1 mg/kg Pb, M/F, n = 191
(96/95)
10 mg/kg Pb, M/F, n = 192
(96/96)
PND 29
Control (vehicle), M/F,
n = 192 (96/96)
1 mg/kg Pb, M/F, n = 192
(96/96)
10 mg/kg Pb, M/F, n = 192
(96/96)
6-61
-------
Study
Species (Stock/Strain), n, Timing of
Sex Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (|jg/dL)a
Endpoints
Examined
Wildemann et al.
(2015)
Rat (Wistar)
Control (vehicle), M, n = 6
NR
Control group provided tap water with
0.2% nitric acid
1.4 ± 1.2 |jg/L for
0 pg/kg/d (0.14 pg/dL)
Spleen weight
357 pg/kg/d Pb, M, n = 5
1607 pg/kg/d Pb, M, n = 5
17 ± 7 pg/L for
357 pg/kg/d (1.77 ± 0.7 pg/dL)
86 ± 29 pg/L for 1607 pg/kg/d
(0.14 pg/dL for 0 pg/kg/d,
1.77 ± 0.7 pg/dL for
357 pg/kg/d, 8.6 ±2.9 pg/dL
for 1607 pg/kg/d)
BLL = blood lead level; d = day; M = male; M/F = male/female; F = female; hr = hour; ISO = isolation, min = minute(s); NR = not reported; Pb = lead; PbO NPs = lead oxide
nanoparticles; PND = postnatal day; ppm = parts per million; w/o = without; wk = week.
alf applicable, reported values for BLL were converted to mg/dL using WebPlot Digitizer (https://apps.automeris.io/wpd/') and are shown in parenthesis.
6-62
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Table 6-11
Animal toxicological studies of white blood cell counts and differentials (spleen, thymus, lymph
node, bone marrow)
Study
Species (Stock/Strain),
n, Sex
Timing of Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (|jg/dL)a
Endpoints
Examined
Cai et al. (2018)
Rat (Sprague Dawley)
Control (vehicle), M/F,
n = 5
0.2% Pb, M/F, n = 5
8-10 wk to 20-30 wk
Rats were 8-10 wk old when
acquired. Whether or not the rats
were allowed to acclimate to the
facility prior to study initiation was
not reported. The number of males
and females not reported.
Control animals received tap water
20.5 ± 0.68 pg/L for 0%
(2.2 ± 6.4 pg/dL)
87.4 ± 9.2 pg/L for 0.2%
(9.3 ± 0.98 pg/dL)
Bone marrow
cell counts
and
differentials
Fanq et al. (2012)
Rat (Sprague Dawley)
Control (vehicle), M,
n =20
300 ppm Pb, M, n = 20
23-25 d to 65-67 d
Dosing solutions were changed
twice per week
4.48 pg/dL for 0 ppm
18.48 pg/dL for 300 ppm - d
65-67
Thymus cell
counts and
differentials,
Spleen cell
counts and
differentials,
Lymph node
cell counts
and
differentials
Yathapu et al. (2020)
Rat (Sprague Dawley)
Control (vehicle), M/F,
n = 32 (16/16)
500 ppm Pb, M/F, n = 32
(16/16)
PND 54-PND 82
Weanling rats (PND 21) were
acclimated to the facility for 5 days
before being divided into two groups
(n = 16) to begin a 28-day long Fe
deficiency diet. After 28 d, the rats
were exposed to Pb or control diet
(n = 16). At this point (PND 82),
blood was collected from rats before
immunization with TT (n = 8)
followed by two boosters
administered in 2 wk intervals.
Vaccine response was evaluated
2 wk later
2.1 ± 1.0 pg/dL for
0 mg/4 mL/kg - PND 82,
Control diet
Spleen cell
counts and
differentials
16.1 ± 5.5 pg/dL for
25 mg/4 mL/kg - PND 82,
Control diet
1.9 ± 0.7 pg/dL for
0 mg/4 mL/kg - PND 82, Iron
deficiency diet
41.6 ± 10.2 pg/dL for
25 mg/4 mL/kg - PND 82,
Iron deficiency diet
6-63
-------
Study
Species (Stock/Strain),
n, Sex
Timing of Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (|jg/dL)a
Endpoints
Examined
Zhu et al. (2020)
Mouse (C57BL.6) 7-9 wk
Control (vehicle), M/F,
n = NR
125 ppm Pb, M/F, n = NR
1250 ppm Pb, M/F,
n = NR
Control animals were exposed to
drinking water containing sodium
acetate. The investigators specified
that an equal number of male and
female mice were used in the study,
but the number of animals used in
some analyses was not an even
number. Consequently, it is not
possible to determine sex
composition ofthe group and it
suggests there may have been
unreported attrition
0 |jg/dL for 0 ppm
4.7 ± 0.2 |jg/dL for 125 ppm
41.3 |jg/dL for 1250 ppm
Spleen cell
counts and
differentials,
Bone marrow
cell counts
and
differentials,
Lymph node
cell counts
BLL = blood lead level; d = day; M/F = male/female; NR = not reported; Pb = lead; PND = postnatal day; ppm = parts per million; wk = week; TT = tetanus toxoid.
alf applicable, reported values for BLL were converted to mg/dL using WebPlot Digitizer (https://apps.automeris.io/wpd/') and are shown in parenthesis.
6-64
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Table 6-12
Animal toxicological studies of white blood cell counts (hematology and subpopulations)
Study
Species (Stock/Strain),
n, Sex
Timing of Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (|jg/dL)a
Endpoints
Examined
Andielkovic et al.
(2019)
Rat (Wistar)
Control (vehicle), M,
n = I
0.2% Pb, M, n = 6
NR
Rats (250 g), age at time of dosing not 24.9 ± 19 |jg/kg for 0 mg WBC counts,
reported, were exposed to a single
dose of 150 mg Pb/kg BW Pb acetate
via oral gavage. Control animals were
given "water"
Pb/kg BW (2.6 ± 2.0 pg/dL) WBC
subpopulations
291.2 ± 139 pg/kg for
150 mg Pb/kg BW
(30.9 ± 14.7 pg/dL)
Cai et al. (2018)
Rat (Sprague Dawley)
Control (vehicle), M/F,
n = 5
0.2% Pb, M/F, n = 5
-10 wk to 20-30 wk
Rats were 8-10 wk old when acquired.
Whether or not the rats were allowed
to acclimate to the facility prior to study
initiation was not reported. The
number of males and females not
reported
Control animals received tap water
20.5 ± 0.68 pg/L for 0%
(2.2 ± 6.4 pg/dL)
87.4 ± 9.2 pg/L for 0.2%
(9.3 ± 0.98 pg/dL)
WBC counts
Corsetti et al. (2017)
Mouse (C57BJ)
Control (vehicle), M,
n = I
30-75 d
200 ppm Pb, M, n = 8
Mice were exposed via drinking water
for 45 consecutive days. Control
animals were exposed to drinking
water containing acetic acid (1 mL/L)
<5 pg/dL for 0 ppm
21.6 pg/dL for 200 ppm
WBC counts
Zhu et al. (2020)
Mouse (C57BL.6) 7-9 wk
Control (vehicle), M/F,
n = NR
125 ppm Pb, M/F, n = NR
1250 ppm Pb, M/F,
n = NR
Control animals were exposed to
drinking water containing sodium
acetate. The investigators specified
that an equal number of male and
female mice were used in the study,
but the number of animals used in
some analyses was not an even
number. Consequently, it is not
possible to determine sex composition
of the group and it suggests there may
have been unreported attrition
0 pg/dL for 0 ppm
4.7 ± 0.2 pg/dL for
125 ppm
41.3 pg/dL for 1250 ppm
WBC
subpopulations
BW = body weight; d = day; F = female; M = male; M/F = male/female; NR = not reported; Pb = lead; ppm = parts per million; WBC = white blood cell; wk = week.
alf applicable, reported values for BLL were converted to mg/dL using WebPlot Digitizer (https://apps.automeris.io/wpd/) and are shown in parenthesis.
6-65
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Table 6-13
Epidemiologic studies of exposure to Pb and sensitization and allergic response
Reference and
Study Design
Study Population
Exposure Assessment
Outcome Confounders
Effect Estimates
and 95% Clsa
tAshlev-Martin et al.
Maternal-Infant
Maternal/Cord Blood
IL-33, TSLP, and IgE Age
ORs per 10-fold
(2015)
Research on
increase in Pb
Environmental
Blood Pb was measured in whole
IL-33, TSLP, and IgE
Canada
Chemicals Study
blood using ICP-MS;
measured in cord blood
Elevated IL-33/TSLP
2008-2011
n: 1256
concentrations measured in the
plasma using a commercial
Cohort
Pregnant women
first and third trimester were
averaged to create an index of
antibody kit and ELISA.
0.72 (0.48, 0.95)
were recruited at
exposure throughout pregnancy
Age at Outcome:
Elevated IgE
<4 wk gestation.
Age at measurement:
At birth
Singleton non-pre-
First and third trimesters
0.98 (0.66, 1.3)
term births
Median: 0.62 pg/dL
Maximum: 4.14 pg/dL
Joseph et al. (2005) n: 4,634
Southeastern
Michigan
1994-1997
Enrollment
(Follow-up for 12 mo
after Pb screening)
Cohort
Children enrolled in
a managed care
organization.
Enrollment at 4 mo
to 3 yr
Blood
Blood Pb measured in venous
whole blood using GFAAS.
Age at measurement: 4 mo to 3 yr
Mean: 4.7 pg/dL (SD: 4.0)
Incident Asthma
Four or more asthma-
medication-dispensing
events in 12 mo or met one
or more of the following
within a 12-mo period:
emergency department visit
for asthma, hospitalization
for asthma, or four or more
outpatient visits for asthma
with at least two asthma-
medication-dispensing
events
Sex, birth weight, and
average annual income
available only at census
block level
HRs:
White children,
>5 vs. <5 pg/dL:
2.7 (0.9, 8.1)
Black children,
>10 vs. <5 pg/dL:
1.3 (0.6, 2.6)
6-66
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
(EEs) and 95% Clsa
tKimetal. (2019)
Cohort for
Childhood Origin of
Maternal/Cord Blood
Atopic Dermatitis and IL-13
Gender and parental
history of allergic diseases
EEs per unit increase
in In(Pb)
Seoul
Asthma and
Cord blood Pb measured using
IL-13 measured in cord
South Korea
Allergic Disease
ICP-MS
blood; diagnosis of atopic
HR Atopic Dermatitis
2007-2011
n: 331
Age at measurement:
dermatitis by pediatric
enrollment (at least
At birth
allergists, and atopic
0.96 (0.60, 1.53)
2 yr of follow-up)
Pregnant women
dermatitis scored using a
Cohort
enrolled in third
Median: 1.3 |jg/dL
validated measure
In(SCORAD)
trimester, children
Maximum: 4.3 |jg/dL
(SCORAD)
followed at least
Atopic Dermatitis
2 yr
Age at Outcome:
At birth (IL-13), 6 mo, 12
mo, and 2 yr
Severity
1.11 (-2.65, 4.87)
IL-13 (pg/ml)
0.69 (0.11, 1.28)
tbKim et al. (2013)
Mothers' and
Children's
Maternal/Cord Blood
Atopic Dermatitis
Age, weight, history of
atopic disease, maternal
OR
South Korea
Environmental
Cord blood Pb measured using
Age at Outcome:
education, infant sex,
Atopic Dermatitis
2006-2009
Health Study
GFAAS
6 mo
family income, family size,
enrollment (follow-up
n: 637
Age at measurement: At birth
parity, duration of breast
1.05 (0.60, 1.81)
with infants 6 mo
feeding, passive smoking
after birth)
Singleton children
Mean: 1.01 pg/dL
during pregnancy, and
Cohort
of mothers enrolled
between weeks 12
and 28 of gestation
cord blood cadmium
tKimetal. (2016)
KNHANES
n: 2184
Blood
igE
Age, sex, urine cotinine,
mercury, and cadmium
% Change in Total
IgE (kU/L)
South Korea
Blood Pb was measured in
Serum total IgE (kU/L)
2010-2011
General population;
venous whole blood using GFAAS
measured by
Sensitization
Negative
Cross-Sectional
26-55 yr old
Age at measurement:
26-55 yr old
immunoradiometric assay
Age at Outcome:
3.5% (-1.8%, 9.4%)
Median: 2.14 |jg/dL
26-55 yr old
Sensitization
75th: 2.82 pg/dL
Positive
10.4% (3.3%, 17.8%)
6-67
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
(EEs) and 95% Clsa
tMeneretal. (2015)
NHANES
Blood
Immune System Effects
Age, sex, ethnicity, BMI,
ORs
n: 2,712 children;
exposure to tobacco
Increased sensitivity
United States
4,333 adults
Blood Pb was measured in
Food Allergen-Specific
smoke, asthma, musty
to food allergens
2005-2006
venous whole blood using ICP-MS
Serum IgE measured using
smell, presence of
Cross-Sectional
General population;
Age at measurement:
immunoassays
cockroaches, and
children 6-19 yr
>6 yr old
domestic animals living at
Children
old, adults >20 yr
Age at Outcome:
home, and year home
0.72 (0.48, 0.95)
old
Serum median:
>6 yr old
was built
Children: 0.87 |jg/dL;
Adults
Adults: 1.48 pg/dL
0.98 (0.66, 1.3)
75th:
Children: 1.31 pg/dL;
Adults: 2.34 pg/dL
6-68
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
(EEs) and 95% Clsa
tPesce etal. (2021)
Nancy and Poitier
France
2003-2006
Enrollment (Follow-
up to 8 yr of age)
Cohort
EDEN Birth Cohort
n: 651
Pregnant women
enrolled early in
pregnancy, children
followed through
8 yr of age
Maternal/Cord Blood
Maternal blood Pb measured
between 24 and 28 gestational
weeks using GFAAS; Cord blood
Pb measured at birth using
GFAAS
Age at measurement:
Prenatal
Mean:
Cord blood: 1.45 |jg/dL; Maternal
blood: 1.91 |jg/dL; Median: Cord
blood: 1.2 pg/dL; Maternal blood:
1.7 Mg/dL
Atopic Diseases
Parental questionnaires
using validated questions
from the International Study
on Asthma and Allergies in
Childhood
Age at Outcome:
4, 8, and 12 mo; 2,
and 5 yr; and 8 yr
3, 4,
Sex, Maternity Center,
BMI, maternal education,
parental smoking,
parental history of allergy,
maternal smoking in
pregnancy, birth weight,
gestational age at
delivery, type of delivery,
manganese, and
cadmium
ORs (Q4, Q1)
Maternal Blood
(>2.2 vs. <1.2):
Asthma
1.25 (0.71, 2.2)
Rhinitis
0.86 (0.51, 1.43)
Eczema
1.04 (0.73, 1.48)
Food Allergy
1.02 (0.51, 2.01)
75th:
Cord blood: 1.8 pg/dL;
Maternal blood: 2.2 |jg/dL
Cord Blood
(>1.8 vs. <0.9):
Asthma
0.74 (0.41, 1.33)
Rhinitis
0.64 (0.37, 1.11)
Eczema
1.35 (0.92, 1.98)
Food Allergy
0.57 (0.25, 1.34)
6-69
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
(EEs) and 95% Clsa
Puah Smith and
Nriaqu (2011)
Saginaw, Ml
Cross-sectional
n: 356
Blood
Children residing in Blood Pb measured in venous
low-income and whole blood
minority
households
identified by the
Statewide Systemic
Tracking of
Elevated Lead
Levels and
Remediation
(STELLAR)
database
Prevalent Asthma
Parental report of asthma
diagnosis
Age, sex, income, stories
in unit, pet ownership,
cockroach problem,
persons in home, smoker
in home, clutter, highest
BLL at address, candles
or incense, months of
residency, housing tenure,
stove type, heating
source, air conditioning
type, peeling paint,
ceiling/wall damage,
housing age, water
dampness
OR
(>10 vs. <10 [jg/dL)
Asthma
7.5 (1.3, 42.9)
Rabinowitz et al.
(1990)
Boston, MA
Enrollment: 1979-
1981
Follow-up: Unclear
Cohort
n: 159
Mother infant pairs
recruited from
Boston Hospital for
Women
Cord Blood
Cord blood Pb measured in
samples at birth using anodic
stripping voltammetry
Prevalent Eczema and
Asthma
Eczema and asthma
prevalence evaluated via
parental questionnaire
N/A
OR
(>10, vs. <10[jg/dL)
Eczema
1.0 (0.6, 1.6)
Asthma
1.3 (0.8, 2.0)
6-70
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
(EEs) and 95% Clsa
tTsuii etal. (2019)
Japan
2011-2014
Cross-sectional
Japan Environment Blood, Maternal/Cord Blood
and Children's
Study
n: 14408
Blood Pb measure using ICP-MS
Age at measurement:
Second/third trimester
Mean: 6.44 ng/g
Q1: <4.79 ng/g
Q4: >7.42 ng/g
Allergen-Specific IgE
Allergen-specific serum IgE
measured using
immunological assays
Age at Outcome:
First trimester
Age, BMI, allergic
diseases, smoking during
pregnancy, smoking
habits of partner, alcohol
consumption during
pregnancy, pet ownership,
month of blood sample,
and geographic region
ORs (Q4, Q1)
House Dust Mite
Sensitization
0.91 (0.83, 1.01)
Japanese Cedar
Pollen Sensitization
1.04 (0.94, 1.15)
Animal Dander
Sensitization
0.99 (0.88, 1.12)
tWeietal. (2019)
United States
2005-2006
Cross-Sectional
NHANES
n: 4509
General population;
all ages
Blood
Eczema
Blood Pb was measured in Self-reported physician's
venous whole blood using ICP-MS diagnosis of eczema
Age at measurement:
>1 yr old Age at Outcome:
>1 yr old
Mean:
Ages >20: 1.75 |jg/dL;
<20: 1.24 |jg/dL
Adults
T1
0.18-1.09
pg/dL
T2
1.09-1.99
pg/dL
T3
2.00-26.4
pg/dL
Children
T1
0.18-0.77
pg/dL
T2
0.78-1.36
pg/dL
T3
1.37-55.3
pg/dL
Age, gender, ethnicity, ORs
education, poverty-income
ratio, smoking, alcohol
use, sleep, and BMI
T1: Reference
Eczema - Adults:
T2
T3
1.14 (0.75, 1.76)
1.09 (0.62, 1.92)
Eczema - Children:
T1
T2
T3
Reference
0.99 (0.62, 1.58)
0.90 (0.60, 1.35)
6-71
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
(EEs) and 95% Clsa
tWells etal. (2014)
United States
2005-2006
Cross-Sectional
NHANES
n: 1788
General population;
children 2-12 yrold
Blood
Blood Pb was measured in
venous whole blood using ICP-MS
Age at measurement:
2-12 yrold
Geometric Mean: 1.13 pg/dL
Immune System Effects
Serum total IgE),
Eosinophils (WBC
differential from complete
blood counts), Asthma
(parental/guardian
reported), Atopy (at least
one specific IgE >0.35
kU/L), Allergies
(parental/guardian
reported)
Age at Outcome:
2-12 yrold
Season, age, sex,
race/ethnicity, parental
education, presence of
smokers in the home,
prenatal smoke exposure,
BMI, presence of
cockroaches in the home,
and avoidance/removal of
pets
ORs
Asthma
1.01 (0.76,
Atopy
1.05 (0.93,
1.35)
1.18)
% Increase
Total IgE (kU/L)
10.3% (3.5%, 17.5%)
Percent Eosinophils
4.6% (2.4%, 6.8%)
BLL = blood lead level; BMI = body mass index; CI = confidence interval; EDEN = Effect of Diet and Exercise on Immunotherapy and the Microbiome; ELISA = enzyme-linked
immunosorbent assay; EEs = effect estimates; GFAAS = graphite furnace atomic absorption spectrometry; HR = hazard ratio; ICP-MS = inductively coupled plasma mass
spectrometry; Ig- = immunoglobulin type; IL- = interleukin type; KNHANES = Korea National Health and Nutrition Examination Survey; In = natural log; mo = month(s); N/A = not
applicable; NHANES = National Health and Nutrition Examination Survey; NR = not reported; OR = odds ratio; Pb = lead; Q = quartile; SCORAD = scoring atopic dermatitis;
SD = standard deviation; SES = socioeconomic status; T# = fertile #; TSLP = thymic stromal lymphopoietin; vs. = versus; WBC = white blood cell; wk = week; yr = year.
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb level or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results
corresponding to a change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized
accordingly.
fStudies published since the 2013 Pb ISA.
6-72
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Table 6-14
Animal toxicological studies of immediate-type hypersensitivity
study species (Stock/Strain), Tlmlng of Exposure Exposure Details
n, Sex a r (Concentration, Duration)
BLL as Reported (|jg/dL)a
Endpoints Examined
Cai et al. Rat (Sprague Dawley)
Control (vehicle), M/F,
n = 5
0.2% Pb, M/F, n = 5
8-10 wk to 20-30 wk
Rats were 8-10 wk old
when acquired. Whether or
not the rats were allowed to
acclimate to the facility prior
to study initiation was not
reported. The number of
males and females not
reported.
Control animals received tap
water
20.5 ± 0.68 |jg/L for 0%
(2.2 ± 6.4 |jg/dL)
87.4 ± 9.2 |jg/L for 0.2%
(9.3 ± 0.98 |jg/dL)
Blood cytokine levels
Fang et al. Rat (Sprague Dawley)
Control (vehicle), M,
n =20
300 ppm Pb, M, n = 20
23-25 d to 65-67 d Dosing solutions were
changed twice per week
4.48 |jg/dL for 0 ppm
18.48 |jg/dL for 300 ppm-d 65-
67
Blood cytokine levels
BLL = blood lead level; d = day(s); M = male; M/F = male/female; ppm = parts per million; wk = week(s).
alf applicable, reported values for BLL were converted to mg/dL using WebPlot Digitizer (https://apps.automeris.io/wpd/') and are shown in parenthesis.
6-73
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Table 6-15 Epidemiologic studies of exposure to Pb and autoimmunity and autoimmune disease
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tJoo etal. (2019)
KNHANES
n: 32215
Blood
Rheumatoid Arthritis
Age, sex, SES, and
smoking status
Rheumatoid Arthritis
(OR)
South Korea
Blood Pb was measured
Self-reported physician
2008-2013
General population
in venous whole blood
diagnosis of rheumatoid
1.01 (0.89, 1.14)
Cross-sectional
using GFAAS
Age at measurement:
All ages
Mean: Rheumatoid
Arthritis: 2.38 |jg/dL;
Control: 2.44 |jg/dL
arthritis
Age at Outcome:
All ages
tKamvcheva et al.
NHANES
Blood
Celiac Disease
Family income to poverty
Celiac Disease
(2017)
n: 3,643 children and 11,040
Seropositivity
ratio and race/ethnicity
Mean difference in BLL
adults
Blood Pb was measured
by CD seropositivity
status
United States
in venous whole blood
Serum tTG-lgA analyzed
2009-2012
General population, >6 yr old
using ICP-MS
with an ELISA
Cross-sectional
Age at measurement:
>6 yr
Age at Outcome:
Adults
>6 yr
-0.17 |jg/dL (-0.54, 0.20)
Mean: Non-Hispanic
white: 1.39 |jg/dL; other
race/ethnicity:
Children
1.47 |jg/dL
-0.14 |jg/dL (-0.27,
-0.02)
BLL = blood lead level; CD = cluster of differentiation; CI = confidence interval; ELISA = enzyme-linked immunosorbent assay; Ig- = immunoglobulin type; GFAAS = graphite furnace
atomic absorption spectrometry; ICP-MS = inductively coupled plasma mass spectrometry; KNHANES = Korea National Health and Nutrition Examination Survey;
NHANES = National Health and Nutrition Examination Survey; OR = odds ratio; Pb = lead; SES = socioeconomic status; tTG-lgA = tissue transglutaminase immunoglobulin A;
yr = year(s)
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb level or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results
corresponding to a change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized
accordingly.
fStudies published since the 2013 Pb ISA.
6-74
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Table 6-16
Animal toxicological studies of autoimmunity and autoimmune disease
Study
Species (Stock/Strain), Tjmjng of Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (|jg/dL)a
Endpoints Examined
Fanq et al.
(2012)
Rat (Sprague Dawley) 23-25 d to 65-67 d
Control (vehicle), M,
n =20
300 ppm Pb, M, n = 20
Dosing solutions were
changed twice per week
4.48 |jg/dL for 0 ppm
18.48 |jg/dL for 300 ppm-
d 65-67
Treg cell suppression assay
BLL = blood lead level; d = day; M = male; ppm = parts per million; Treg = regulatory T cell.
alf applicable, reported values for BLL were converted to mg/dL using WebPlot Digitizer (https://apps.automeris.io/wpd/') and are shown in parenthesis.
6-75
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_ EPA/600/R-23/375
£% United States
Environmental Protection January 2024
m m Agency www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 7: Hematological Effects
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
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CONTENTS
DOCUMENT GUIDE 7-iii
LIST OF TABLES 7-v
LIST OF FIGURES 7-vi
ACRONYMS AND ABBREVIATIONS 7-vii
APPENDIX 7 HEMATOLOGICAL EFFECTS 7-1
7.1 Introduction, Summary of the 2013 Integrated Science Assessment for Lead, and Scope
of the Current Review 7-1
7.1.1. Red Blood Cell Survival and Function 7-2
7.1.2. Heme Synthesis 7-2
7.2 Scope 7-3
7.3 Red Blood Cell Survival and Function 7-5
7.3.1. Epidemiologic Studies of Red Blood Cell Survival and Function 7-5
7.3.2. Toxicological Studies of Red Blood Cell Survival and Function 7-7
7.3.3. Integrated Summary of Red Blood Cell Survival and Function 7-9
7.4 Heme Synthesis 7-10
7.4.1. Epidemiologic Studies of Heme Synthesis 7-10
7.4.2. Toxicological Studies of Heme Synthesis 7-10
7.4.3. Integrated Summary of Heme Synthesis 7-11
7.5 Biological Plausibility 7-11
7.5.1. Decreased Red Blood Cell Survival and Function 7-14
7.5.2. Altered Heme Synthesis 7-15
7.6 Summary and Causality Determination 7-17
7.6.1. Causality Determination for Red Blood Cell Survival and Function 7-17
7.6.2. Evidence for Red Blood Cell Survival and Function 7-17
7.6.3. Evidence for Heme Synthesis 7-18
7.6.4. Causality Determination 7-19
7.7 Evidence Inventories—Data Tables to Summarize Study Details 7-25
7.8 References 7-32
7-iv
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LIST OF TABLES
Table 7-1 Summary of evidence indicating a causal relationship between Pb exposure and
hematological effects 7-20
Table 7-2 Epidemiologic studies of exposure to Pb and hematological effects 7-25
Table 7-3 Animal toxicological studies of Pb exposure and hematological effects 7-30
7-v
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LIST OF FIGURES
Figure 7-1 Potential biological plausibility pathways for hematological effects associated with
exposure to Pb. 7-13
7-vi
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ACRONYMS AND ABBREVIATIONS
ALAD S-aminolevulinic acid dehydratase
ALA aminolevulinic acid
AQCD Air Quality Criteria Document
BLL blood lead level
BMI body mass index
BW body weight
Ca2+ calcium ion(s)
CAT catalase
CI confidence interval
e-waste electronic waste
EPO erythropoietin
Fe2+ iron ion
GFAAS graphite furnace atomic absorption
spectrometry
GPx glutathione peroxidase
GR glutathione reductase
GSH glutathione
GSSG glutathione disulfide
Hb hemoglobin
Hct hematocrit
HSC hematopoietic stem cell
hr hour(s)
H2O2 hydrogen peroxide
ICP-MS inductively coupled plasma mass
spectrometry
ISA Integrated Science Assessment
KNHANES Korea National Health and Nutrition
Examination Survey
In natural log
M male
M/F male/female
MCH mean corpuscular hemoglobin
MCHC mean corpuscular hemoglobin
concentration
MC V mean corpuscular volume
MDA malondialdehyde
Mg2+ magnesium ion
mo month(s)
OH hydroxyl radical
NCE normochromatic erythrocytes
NR not reported
O2- superoxide
OR odds ratio
Pb lead
PCE polychromatic erythrocytes
PCV packed cell volume
PECOS Population, Exposure, Comparison,
Outcome, and Study Design
Pit platelet
PND postnatal day
PS phosphatidylserine
Q quartile
RBC red blood cell
RDW red blood cell distribution width
ROS reactive oxygen species
SD standard deviation
SES socioeconomic status
SOD superoxide dismutase
U.S. EPA United States Environmental Protection
Agency
wk week(s)
yr year(s)
Zn zinc
ZPP zinc-protoporphyrin
7-vii
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APPENDIX 7
HEMATOLOGICAL EFFECTS
Causality Determination for Pb Exposure and Hematological Effects
This appendix characterizes the scientific evidence that supports the causality
determination for lead (Pb) exposure and hematological effects. The types of studies
evaluated within this appendix are consistent with the overall scope of the ISA as
detailed in the Process Appendix (see Section 12.4). In assessing the overall evidence,
the strengths and limitations of individual studies were evaluated based on scientific
considerations detailed in Table 12-5 of the Process Appendix (Section 12.6.1). More
details on the causal framework used to reach these conclusions are included in the
Preamble to the ISA (U.S. EPA. 2015). The evidence presented throughout this
appendix supports the following causality determination:
Outcome Causality Determination
Hematological Effects, Including Altered
Heme Synthesis and Decreased Red Causal
Blood Cell Survival and Function
The Executive Summary, Integrated Synthesis, and all other appendices of this Pb
ISA can be found at https://assessments.epa.gov/isa/document/&deid=359536.
7.1 Introduction, Summary of the 2013 Integrated Science
Assessment for Lead, and Scope of the Current Review
Hematology is a subgroup of clinical pathology concerned with the morphology, physiology, and
pathology of blood and blood-forming tissues. Hematological measures, when evaluated with information
on other biomarkers, are informative diagnostic tests for blood-forming tissues (i.e., bone marrow, spleen,
and liver) and organ function.
The 2013 Integrated Science Assessment for Lead (hereinafter referred to as the 2013 Pb ISA)
issued causality determinations for hematological effects resulting from lead (Pb) exposure on red blood
cell (RBC) survival and function and altered heme synthesis (U.S. EPA. 2013). The evidence
underpinning these causality determinations is summarized below. Given the interconnectedness of the
effects of Pb on RBC survival and function and altered heme synthesis, this assessment presents a single
causality determination for Pb exposure and hematological effects. This approach allows for a more
holistic evaluation of inter-related health endpoints, including a discussion of how all individual lines of
evidence contribute to the overall hematological effects causality determination.
7-1
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7.1.1.
Red Blood Cell Survival and Function
The body of epidemiologic and toxicological evidence assessed in the 2013 Pb ISA indicates a
"causal" relationship between Pb exposure and decreased RBC survival and function. Experimental
animal studies demonstrate that relevant human blood Pb levels (BLLs) from oral and inhalation exposure
alter several hematological parameters (e.g., RBC number), increase measures of oxidative stress
(e.g., inhibition of antioxidant enzymes in RBCs), and increase cytotoxicity in RBC precursor cells. Some
of these effects have been observed in animal toxicological studies with exposures resulting in BLLs of
2-7 (ig/dL. Evidence of biologically plausible modes of action, including increased intracellular calcium
2+
concentrations [Ca ] decreased Ca2+/Mg2+ ATPase activity, and increased phosphatidylserine (PS)
exposure leading to RBC destruction by macrophages, support these findings. Epidemiologic studies
reported associations between exposure to Pb, BLL, and altered hematological endpoints, increased
measures of oxidative stress, and altered hematopoiesis in adults and children. Although most of these
studies are limited by their lack of rigorous methodology (i.e., correlations, t tests, or chi squared
analyses), some studies in children did adjust for potential confounding factors, including age, sex,
mouthing behavior, anemia, dairy product consumption, maternal age, education, employment, marital
status, family structure, and socioeconomic status (SES)-related variables. Though limited in number,
studies that adjusted for confounders also reported consistent associations between BLL and altered
hematological parameters, strengthening their support for findings in experimental animals. Collectively,
the strong evidence from toxicological studies supported by findings from mode of action and
epidemiologic studies reviewed in the 2013 Pb ISA was sufficient to conclude that there is a causal
relationship between Pb exposures and decreased RBC survival and function.
7.1.2. Heme Synthesis
Available toxicological evidence evaluated in the 2013 Pb ISA indicated a causal relationship
between Pb exposure and altered heme synthesis (U.S. EPA. 2013). Altered heme synthesis is
demonstrated by a small but consistent body of studies in adult animals, which report that exposures that
result in BLLs relevant to humans (e.g., 10 (ig/dL) lead to decreased S-aminolevulinic acid dehydratase
(ALAD) and ferrochelatase activities. Supporting this evidence is a larger body of ecotoxicological
studies that demonstrate decreased ALAD activity across a wide range of taxa exposed to Pb. Evidence of
biologically plausible modes of action, including evidence that Pb acts directly on two enzymes involved
in heme synthesis (ALAD and ferrochelatase), decreased RBC Hb concentration, measures of oxidative
stress, and evidence that administration of antioxidants reduced the effects of Pb exposure on antioxidant
enzymes, support these findings. Epidemiologic studies find associations in both adults and children
between higher BLLs and decreased ALAD and ferrochelatase activities. Although most of these studies
are limited by their lack of rigorous methodology and consideration of potential confounders, some
studies in children did incorporate potential confounding factors (i.e., age, sex, urban/rural residence,
7-2
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height, weight, body mass index [BMI]). Although limited in number, studies that adjusted for
confounders also reported consistent associations between BLLs and decreased ALAD and ferrochelatase,
strengthening support for the findings in the animal toxicological studies. Evidence for altered heme
synthesis is also provided by a large body of toxicological and epidemiologic studies that report decreased
hemoglobin (Hb) concentrations in association with Pb exposure or BLL. The 2013 Pb ISA concluded
that, collectively, the strong evidence from toxicological and ecotoxicological studies—supported by
findings from epidemiologic studies—is sufficient to conclude a causal relationship between Pb
exposures and altered heme synthesis.
This ISA determined causality for adverse effects of Pb exposure on altered heme Synthesis and
decreased RBC survival and function. Recent evidence demonstrates that Pb exposures alter several
hematological parameters, decrease enzyme activity related to heme synthesis, and increase RBC
oxidative stress. Biological plausibility is provided by toxicological and epidemiologic studies
demonstrating increased intracellular calcium concentrations, decreased Ca2+/Mg2+ ATPase activity, and
increased PS exposure. Taken together, there is sufficient evidence to conclude that there is a causal
relationship between Pb exposure and hematological effects, including altered heme synthesis and
decreased RBC survival and function.
The following sections provide an overview of the scope of the appendix (Section 7.2), evaluation
of the scientific evidence relating Pb exposures and hematological effects (Sections 7.3 and 7.4), a
discussion of biological plausibility (Section 7.5), and a summary section with the updated causality
determination (Section 7.6, Table 7-1). The focus of these sections is on studies published since the
completion of the 2013 Pb ISA (U.S. EPA. 2013). Study-specific details, including animal type, exposure
concentrations and exposure durations in experimental studies, and study design; exposure metrics; and
select results in epidemiologic studies are presented in evidence inventories in Section 7.7.
7.2 Scope
The scope of this section is defined by Population, Exposure, Comparison, Outcome, and Study
Design (PECOS) statements. The PECOS statement defines the objectives of the review and establishes
study inclusion criteria, thereby facilitating identification of the most relevant literature to inform the Pb
ISA.1 To identify the most relevant literature, the body of evidence from the 2013 Pb ISA was considered
in the development of the PECOS statements for this appendix. Specifically, well-established areas of
research; gaps in the literature; and inherent uncertainties in specific populations, exposure metrics,
'The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
7-3
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comparison groups, and study designs identified in the 2013 Pb ISA inform the scope of this appendix.
The 2013 Pb ISA used different inclusion criteria than the current ISA, and the studies referenced therein
often do not meet the current PECOS criteria (e.g., due to higher or unreported biomarker levels). Studies
that were included in the 2013 Pb ISA, including many that do not meet the current PECOS criteria, are
discussed in this appendix to establish the state of the evidence prior to this assessment. With the
exception of supporting evidence used to demonstrate the biological plausibility of Pb-associated
hematological effects, recent studies were only included if they satisfied all of the components of the
following discipline-specific PECOS statements:
Epidemiologic Studies
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect;
Exposure: Exposure to Pb2 as indicated by biological measurements of Pb in the body, with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure3; or intervention groups in randomized trials and quasi-experimental studies;
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles);
Outcome: Hematological effects including but not limited to disruption of heme synthesis and
RBC survival and function; and
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials, and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages);
2Recent studies of occupational exposure to Pb were only considered insofar as they addressed a topic area that was
relevant to the National Ambient Air Quality Standards review (e.g., longitudinal studies designed to examine recent
versus historical Pb exposure).
3Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 |im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 |im3 (PM2 5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are limited, it is difficult to
assess the representativeness of these concentrations to population exposure [Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PNfc.s with BLLs are lacking.
7-4
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Exposure: Oral, inhalation or intravenous treatment(s) administered to a whole animal (in
vivo) that results in a BLL of 30 pg/dL or below;4'5
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control;
Outcome: Hematological effects; and
Study design: Controlled exposure studies of animals in vivo.
7.3 Red Blood Cell Survival and Function
Toxicological and epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA, 2013)
provided strong evidence that exposure to Pb affects a range of hematological outcomes related to RBC
survival and function; which is consistent with epidemiologic evidence from the 2006 Pb Air Quality
Criteria Document (AQCD) (U.S. EPA, 2006), demonstrating an association between high BLLs and
anemia in children. Given the extensive evidence base at higher BLLs, the scope for this appendix focuses
on toxicological studies conducted at lower exposure levels and epidemiologic studies in nonoccupational
populations (as described in Section 7.2). Under the defined PECOS criteria, recent toxicological and
epidemiologic studies provide additional support for Pb-related changes in Hb concentration and some
other hematological measures of RBC survival and function. Below, recent evidence is reviewed in the
context of evidence from past assessments.
7.3.1. Epidemiologic Studies of Red Blood Cell Survival and Function
The epidemiologic evidence evaluated in the 2013 Pb ISA (U.S. EPA. 2013) covered a range of
measures related to RBC survival and function, including RBC counts and other hematological
parameters, hematopoiesis, Ca2+/Mg2+ ATPase activity, PS exposure, and RBC oxidative stress.
Specifically, epidemiologic studies provided evidence that elevated BLLs in children and adults are
associated with altered hematological parameters (e.g., decreased RBC counts, Hb concentration, and
hematocrit [Hct] and changes in mean corpuscular volume [MCV] and mean corpuscular hemoglobin
[MCH]), increased measures of oxidative stress (e.g., altered antioxidant enzyme activities [superoxide
dismutase (SOD), catalase (CAT), glutathione peroxidase (GPx)], decreased cellular glutathione (GSH),
and increased lipid peroxidation), and altered hematopoiesis (e.g., decreased erythropoietin [EPO]).
Notably, most of these epidemiologic studies are cross-sectional in design and conducted either in
4Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
5This level is approximately an order of magnitude above the upper end of the distribution of U.S. young children's
BLLs. The 95th percentile of the 2011-2016 National Health and Nutrition Examination Survey distribution of BLL
in children (1-5 years; n = 2,321) is 2.66 (ig/dL dL (CDC. 2019) and the proportion of individuals with BLL that
exceed this concentration varies depending on factors including (but not limited to) housing age, geographic region,
and a child's age, sex and nutritional status.
7-5
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occupationally exposed populations or other populations with higher mean Pb exposures (i.e., BLLs
>10 (ig/dL). These studies were additionally limited by their lack of consideration of potential
confounders, although some studies in children did adjust for a range of factors, including age, sex,
mouthing behavior, anemia, dairy product consumption, maternal age, education, employment, marital
status, family structure, and SES-related variables (Oucirolo et al.. 2010; Ahamed et al.. 2007; Riddell et
al.. 2007). Studies that did account for potential confounders reported consistent associations between
higher BLLs and lower Hb levels, higher prevalence of anemia, and higher levels of RBC oxidative stress
(Sections 4.7.2.1 and 4.7.2.7 of the 2013 Pb ISA). As a whole, the cross-sectional study designs, higher
exposures, and lack of rigorous statistical methodologies in many studies raise uncertainties in the
epidemiologic evidence regarding the directionality of effects; the level, timing, frequency, and duration
of Pb exposure that contributed to the observed associations; and whether the observed associations are
independent of potential confounders.
Recent epidemiologic studies provide generally consistent evidence of inverse associations
between Pb exposures and Hb levels in children. These associations are reported at lower BLLs than in
studies included in the 2013 Pb ISA. Evidence for associations with other hematological parameters of
RBC function and survival, as well as Hb levels in adults, is less robust. Consistent with the 2013 Pb ISA,
most recent studies are cross-sectional analyses, which are unable to establish temporality between
exposure and outcome. Recent studies include populations with lower mean BLLs and more robust
adjustment for potential confounders compared with studies included in the 2013 Pb ISA. Measures of
central tendency for BLLs used in each study, along with other study-specific details including study
population characteristics and select effect estimates, are highlighted in Table 7-2. An overview of the
recent evidence is provided below.
The most common hematological parameter evaluated in recent epidemiologic studies is Hb
levels. A number of cross-sectional studies of children in China observed lower Hb levels associated with
higher blood Pb or erythrocyte Pb levels (Guo et al.. 2021; Kuang et al.. 2020; Li et al.. 2018; Liu et al..
2015; Liu et al.. 2012). Kuang et al. (2020) and Li et al. (2018) also reported inverse associations between
BLLs and MCH in children, which is a measure of average Hb concentration in a single erythrocyte.
Notably, all of these studies had mean and/or median BLLs below 10 (ig/dL, including some below
5 (ig/dL (Guo et al.. 2021; Kuang et al.. 2020; Liu et al.. 2012). While only a few studies attempted to
account for potential confounding by SES (Kuang et al.. 2020; Liu et al.. 2015). others adjusted directly
for iron deficiency (Li et al.. 2018; Liu et al.. 2012). which may be the direct mechanism by which SES
could potentially confound the relationship between BLLs and Hb (i.e., via nutritional deficiency).
Although the magnitude of the observed effect estimates was not directly comparable across studies,
BLLs were negatively associated with Hb levels in all studies. Notably, Liu et al. (2012) reported that
each 1 (ig/dL higher BLL was associated with a larger decrement in Hb levels when restricting their
sample to children with BLLs less than 10 (ig/dL (-0.174 g/dL [95% confidence interval (CI): -0.27,
-0.078 g/dL]) compared with the full sample (-0.096 g/dL [95% CI: -0.18, -0.012 g/dL]). In studies that
examined quantiles of exposure, blood or erythrocyte Pb levels were only associated with Hb levels at the
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higher quantiles (Guo et al.. 2021; Liu et al.. 2015). For example, in a large hospital-based study in China,
children with BLLs between 3.33 and 4.50 (ig/dL (quintile 4) and those with levels greater than
4.50 (ig/dL (quintile 5) had lower Hb levels relative to children with BLLs less than 1.61 (ig/dL
(-0.49 g/L [95% CI: -0.94, -0.04 g/L] and -1.25 g/L [95% CI: -1.71, -0.78 g/L], respectively);
however, null associations were reported for quintiles 2 (1.61-2.44 (ig/dL) and 3 (2.44-3.33 (ig/dL)
relative to the lowest quintile of exposure (Guo et al.. 2021). While the clinical relevance of small mean
decrements in Hb across exposure quintiles is unclear, Guo et al. (2021) also reported that the odds of
anemia (defined as Hb levels below 110 g/L) were monotonically higher in association with higher blood
Pb quintiles, with 45% (95% CI: 26%, 67%) higher odds of anemia for children in the highest quintile of
exposure relative to the lowest quintile. Similarly, Li et al. (2018) observed 5% (95% CI: 0%, 11%)
higher odds of low Hb levels (<115 g/L) per 1 (ig/dL higher BLL.
Recent studies of Pb exposure and Hb levels in adults are more limited in number and include
overlapping study populations. In contrast to studies in children, Park and Lee (2013) reported higher Hb
associated with higher BLLs for adult participants of the 2008—2010 cycles of the Korea National Health
and Nutrition Examination Survey (KNHANES). The analysis was stratified to examine effect
modification by sex and the observed associations were comparable for men and women. A similar study
analyzed the same KNHANES cycles, but examined the population as a whole, rather than stratified by
sex, and also categorized exposure into quartiles (Kim and Lee. 2013). The authors noted similar positive
associations between blood Pb and Hb levels. However, Kim and Lee (2013) observed inverse
associations across exposure quartiles when correcting BLLs for Hct in order to estimate erythrocyte Pb.
As described in the 2006 Pb AQCD (U.S. EPA. 2006). Pb exposure decreases Hct and MCV, meaning
the negative effects of Pb can potentially decrease Pb levels in whole blood.
In addition to studies examining the relationship between Pb exposure and Hb levels, a few recent
cross-sectional studies evaluate associations between BLLs and other hematological parameters. In a
group of children in China, including some living near a battery plant or a lead/-zinc mine, Li et al. (2018)
reported higher odds of low RBC counts and low blood platelets (Pits) in association with higher BLLs.
In contrast, Kuang et al. (2020) observed a positive association with increased RBC counts in a
convenience sample of slightly older boys in Nanjing, China, with notably lower median BLLs
(2.61 (ig/dL compared with 8.38 (ig/dL). In an adult population, a cohort of pregnant women in Durango,
Mexico, La-Llave-Leon et al. (2015) also observed a positive cross-sectional association between RBC
counts and BLLs. Given this small body of studies that examine diverse populations with varying BLLs,
it is difficult to discern methodological or demographic factors contributing to the inconsistent results.
7.3.2. Toxicological Studies of Red Blood Cell Survival and Function
As previously reported in the 2013 Pb ISA, epidemiologic evidence is coherent with experimental
animal studies demonstrating that exposures via drinking water and oral gavage resulting in BLLs
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relevant to what was found in humans affect multiple hematological outcomes related to RBC survival
and function (U.S. EPA. 2013). Specifically, exposure to Pb has been shown to decrease RBC survival,
either through direct effects on RBC membranes leading to increased fragility or through the induction of
eryptosis and eventual phagocytosis by macrophages (U.S. EPA. 2013). Some of these effects have been
observed in animal toxicological studies with exposures resulting in 2-7 (ig/dL BLL. For example, Hb
concentrations in plasma significantly decreased in male mice exposed to Pb nitrate (50 mg/kg BW in
drinking water for 40 days; BLL: 1.72 ± 0.02 (ig/dL) (Sharma et al.. 2010). BLLs >100 (ig/dL were also
associated with decreased RBC survival in laboratory animals.
The evidence is limited and conflicting for the observed effects of Pb exposure on hematopoiesis
in rats and mice. For example, administration of Pb acetate (140, 250, or 500 mg/kg) via oral gavage once
per week for 10 weeks decreased the number of polychromatic RBCs (PCE) and increased numbers of
micronucleated PCEs in female rats (Celik et al.. 2005). Increased micronucleated PCEs were reported in
female and male rats exposed to Pb acetate in drinking water for 125 days, but decreased
PCEs/normochromatic RBCs (NCEs) ratio was only observed in male rats (Alghazal et al.. 2008).
However, in mice exposed to Pb acetate (1 g/L in drinking water for 90 days), PCE increased, but
PCE/NCE was unaffected (Marques et al.. 2006).
In addition, the 2013 Pb ISA reported that Pb exposure significantly decreases several
hematological parameters. In studies reporting BLL relevant to this ISA, decreased RBC counts
(Andielkovic et al.. 2019; Cai et al.. 2018; Sharma et al.. 2010). Hb concentration (Andielkovic et al..
2019; Cai et al.. 2018; Berrahal et al.. 2011; Sharma et al.. 2010; Baranowska-Bosiacka et al.. 2009;
Masso-Gonzalez and Antonio-Garcia. 2009). and Hct (Andielkovic et al.. 2019; Masso-Gonzalez and
Antonio-Garcia. 2009; Masso et al.. 2007) were reported in laboratory studies conducted in rats. In other
studies in which BLLs were not reported, decreased RBC counts (Simsek et al.. 2009; Marques et al..
2006; Lee et al.. 2005). Hb (Wang et al.. 2010b; Simsek et al.. 2009; Lee et al.. 2005). Hct (Molina et al..
2011; Marques et al.. 2006; Lee et al.. 2005). MCV (Wang et al.. 2010b). MCH (Wang et al.. 2010b;
Simsek et al.. 2009). and mean corpuscular hemoglobin concentration (MCHC) (Wang et al.. 2010b;
Simsek et al.. 2009) were reported in laboratory studies conducted in rats and mice. Some toxicological
studies found no evidence of hematological effects (Gautam and Flora. 2010; Lee et al.. 2006).
Recent studies also report effects of Pb exposure on hematological parameters at BLLs relevant to
this ISA. Administration of Pb acetate in drinking water (0.2%; BLL = 9.3 ± 0.98 (ig/dL) for 84 days
resulted in RBC hemolysis, and significantly decreased RBC lifespan and number, and Hb levels, but had
no effect on Pit number in blood collected from Sprague Dawley rats (Cai et al.. 2018). In a different
study, administration of Pb acetate in drinking water (0.150 mg/kg; BLL = 14.7 (ig/dL) for 1 day
decreased RBCs, Pits, Hb concentration, and Hct, but had no effect on MCV, MCH an MCHC in whole
blood collected from adult male Wistar rats (Andielkovic et al.. 2019). Pb acetate treatment increased Hct
and RBC distribution width, decreased MCHC, and had no effect on RBC number, Hb, MCV, and MCH
in adult male C57BJ mice exposed via drinking water (200 ppm; BLL = 21.6 (ig/dL) for 45 days (Corsetti
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et al.. 2017). The potential effects of Pb exposure in rats were also investigated through a combination of
lactational and drinking water exposures. Beginning on postnatal day (PND) 1, dams were given drinking
water containing Pb acetate (50 mg/L). On PND 21, male pups were subsequently administered Pb
acetate (50 mg/L) in drinking water for an additional 40 or 65 days, at which time they were sacrificed.
Hct was significantly reduced in mice exposed until PND 40 (BLL = 12.67 ± 1.68 (ig/dL) whereas Hct
and Hb were significantly decreased at PND 65 (BLL = 7.49 ± 0.78 (ig/dL) (Berrahal et al.. 2011). Study-
specific details, including animal species, strain, sex, and BLLs are highlighted in Table 7-3.
7.3.3. Integrated Summary of Red Blood Cell Survival and Function
Experimental animal studies evaluated in the 2013 Pb ISA (U.S. EPA. 2013) demonstrate that Pb
exposures resulting in BLLs relevant to humans (i.e., <10 (ig/dL) alter several hematological parameters,
increase measures of oxidative stress, and increase cytotoxicity in RBC precursor cells. While
epidemiologic evidence synthesized in the last review was generally coherent with results from the animal
studies, most of the epidemiologic studies evaluated are cross-sectional, were conducted in populations
with higher mean Pb exposures (i.e., BLLs >10 (.ig/dL). did not thoroughly consider potential
confounders, and lacked rigorous statistical methodology. As a result, there were considerable
uncertainties in the epidemiologic evidence regarding the directionality of effects; the level, timing,
frequency, and duration of Pb exposure that contributed to the observed associations; and whether the
observed associations are independent of potential confounders.
Though limited in number, recent PECOS-relevant animal toxicological studies continue to
support the findings from the last review. Specifically, these studies consistently report the effects of Pb
on hematological parameters, including mostly consistent evidence of a Pb-related decrease in Hb. Recent
epidemiologic studies expand on the evidence presented in the 2013 Pb ISA and provide additional
support for the experimental evidence. Although the recent studies are also cross-sectional, they include
populations with much lower BLL means (<10 (ig/dL) and include more robust adjustment for potential
confounding, addressing important uncertainties from the last review. The most consistent epidemiologic
evidence indicates an association between BLLs and decreased Hb levels in children, which is coherent
with the evidence from recent experimental animal studies. While the clinical relevance of small mean
decrements in Hb is unclear, a few of the recent epidemiologic studies include analyses linking increased
BLLs to increased prevalence of anemia. Recent epidemiologic studies of Hb levels in adults were more
limited in number and less consistent than those in children. Additionally, of the relatively few studies
examining RBC counts, the results were also inconsistent.
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7.4 Heme Synthesis
Toxicological and ecotoxicological studies evaluated in the 2006 Pb AQCD (U.S. EPA, 2006)
and the 2013 Pb ISA (U.S. EPA. 2013) provided strong evidence that exposure to Pb affects heme
synthesis. A limited number of epidemiologic studies contributed compelling supporting evidence. Given
the extensive evidence base at higher BLLs, the scope for this appendix focuses on toxicological studies
conducted at lower exposure levels and epidemiologic studies in nonoccupational populations (as
described in Section 7.2). Below, recent evidence is reviewed in the context of evidence from past
assessments.
7.4.1. Epidemiologic Studies of Heme Synthesis
The epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA. 2013) provided evidence that
BLLs in children and adults are associated with decreased activity of enzymes involved in the heme
synthesis pathway, including ALAD and ferrochelatase. Similar to studies of RBC function and survival,
most studies on heme synthesis are cross-sectional in design and conducted either in occupationally
exposed populations and/or in populations with higher mean Pb exposures (i.e., BLL >20 (ig/dL). These
studies were additionally limited by their lack of consideration of potential confounders, although some
studies adjusted for or considered potential confounding factors (i.e., age, sex, urban/rural residence,
height, weight, and BMI) (Wang et al.. 2010a; Ahamed et al.. 2007; Ahamed et al.. 2006). Studies that did
account for potential confounders reported consistent inverse associations between BLLs and ALAD
activity (Section 4.7.3.1 of the 2013 Pb ISA). Evidence for altered heme synthesis is also provided by the
epidemiologic studies discussed in Section 7.3.1 that report lower Hb concentrations in association with
higher Pb exposure or BLLs. Decreased RBC survival and hematopoiesis can be expected to occur
simultaneously, and any effect on Hb levels is likely a combination of the two processes. Outside of the
recent studies on Hb concentrations discussed in Section 7.3.1, there are no recent PECOS-relevant
epidemiologic studies examining the relationship between Pb exposure and heme synthesis.
7.4.2. Toxicological Studies of Heme Synthesis
Pb-induced alterations in heme synthesis occurring at BLLs relevant to this ISA (e.g., 10 (ig/dL)
have been demonstrated convincingly by a small but consistent body of evidence. In brief, Pb exposure in
rats inhibits several enzymes involved in heme synthesis, most notably ALAD, the enzyme that catalyzes
the second, rate-limiting step in heme biosynthesis (Rendon-Ramirez et al.. 2007; Teravama et al.. 1986).
Pb exposure has also been shown to inhibit ferrochelatase, a mitochondrial iron (Fe)-sulfur (S) containing
enzyme that incorporates Fe2+into protoporphyrin IX to create heme (Rendon-Ramirez et al.. 2007).
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Toxicological studies have found that Pb exposures result in increases in markers of oxidative
stress. For example, Lee et al. (2005) reported increased RBC malondialdehyde (MDA), SOD and CAT
levels accompanied by significant decreases in GSH and GPx in rats exposed to Pb (25 mg/kg) once a
week for 4 weeks. In a second drinking water study performed in rats, administration of Pb acetate
(750 mg/kg in drinking water for 11 weeks) resulted in decreased concentrations of plasma Vitamin C,
Vitamin E, nonprotein thiol, and RBC-GSH, with simultaneous increased activity of SOD and GPx
(Kharoubi et al.. 2008). Effects on measures of oxidative stress were also observed in in vitro studies
including increased MDA and decreased SOD and CAT in RBCs (Ciubar et al.. 2007). and decreased
glutathione reductase (GR) activity in human RBCs (Coban et al.. 2007). and decreased GSH and
increased glutathione disulfide (GSSG).
There were no recent toxicology studies investigating the effects of Pb exposure on heme
synthesis that satisfied the PECOS criteria described in Section 7.2 available for this review.
7.4.3. Integrated Summary of Heme Synthesis
A small number of animal toxicological studies evaluated in the 2013 Pb ISA provide consistent
evidence that Pb exposures affect heme synthesis, including Pb-induced decreases in ALAD (Rendon-
Ramirez et al„ 2007; Terayama et al.. 1986) and ferrochelatase activities (Rendon-Ramirez et al.. 2007).
The toxicological evidence was supported by a larger body of ecotoxicological studies that demonstrate
ALAD inhibition in Pb-exposed aquatic and terrestrial invertebrates and vertebrates (Sections 6.3.4.3,
6.4.5.2, 6.4.5.3, and 6.4.15.2 of the 2013 Pb ISA). Ecological evidence from previous reviews
consistently observed Pb-induced ALAD inhibition in multiple species, including birds and fish (U.S.
EPA, 2013, 2006). Some cross-sectional epidemiologic studies evaluated in previous reviews provide
supporting evidence that concurrent BLLs are associated with decreased ALAD and ferrochelatase
activities in both adults and children. The majority of these studies, however, are limited by the lack of
rigorous methodology and consideration of potential confounding.
Recent evidence provided by epidemiologic and toxicological studies of Pb exposure and Hb
levels provides additional support for Pb-related impairment of heme synthesis. These studies are
discussed in more detail in Section 7.3.
7.5 Biological Plausibility
This section describes biological pathways that potentially underlie effects on hematology
measures resulting from exposure to Pb. Figure 7-1 depicts the proposed pathways as a continuum of
upstream events connected by arrows that may lead to downstream events observed in epidemiologic
studies. Evidence supporting these proposed pathways was derived from Sections 7.3 and 7.4 of this ISA,
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evidence reviewed in the 2013 Pb ISA (U.S. EPA. 2013). and recent evidence collected from studies that
may not meet the current PECOS criteria but contain mechanistic information supporting these pathways.
Discussion of how exposure to Pb may lead to hematological effects contributes to an understanding of
the biological plausibility of epidemiologic results. Note that the structure of the biological plausibility
section and the role of biological plausibility in contributing to the weight-of-evidence analysis used in
the 2013 Pb ISA are discussed below.
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Decreased Ca2+
/Mg2+ ATPase
activity
Altered hematopoiesis
ALAD = 5-aminolevulinic acid dehydratase; Ca2+ = calcium ion; Mg2+ = magnesium ion; RBC = red blood cell.
Note: The boxes represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and the arrows indicate a proposed relationship between
those effects. Solid arrows denote evidence of essentiality as provided, for example, by an inhibitor of the pathway used in an experimental study involving Pb exposure. Dotted arrows
denote a possible relationship between effects. Shading around multiple boxes is used to denote a grouping of these effects. Arrows may connect individual boxes, groupings of
boxes, and individual boxes within groupings of boxes. Progression of effects is generally depicted from left to right and color coded (white, exposure; green, initial effect; blue,
intermediate effect; orange, effect at the population level or a key clinical effect). Here, population-level effects generally reflect results of epidemiologic studies. When there are gaps
in the evidence, there are complementary gaps in the figure and the accompanying text below. The structure of the biological plausibility sections and the role of biological plausibility
in contributing to the weight-of-evidence analysis used in the 2022 Pb ISA are discussed in Section 7.6.
Figure 7-1 Potential biological plausibility pathways for hematological effects associated with exposure to
Pb.
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Careful review of the available evidence indicates that exposure to Pb has the potential to
modulate the hematological parameters leading to decreased RBC survival and function and altered heme
synthesis. These deficits converge, promoting the development of anemia, a condition that occurs when
the number of RBCs and/or the concentration of Hb in RBCs is abnormally low. Below, evidence from
peer-reviewed toxicology studies providing biological plausibility for Pb-associated effects on
hematological parameters is reviewed.
7.5.1. Decreased Red Blood Cell Survival and Function
As described below, there is strong evidence that Pb impacts a series of hematological parameters
along a cascade of events that results in decreased RBC function and survival, possibly leading to anemia
(Figure 7-1). As reviewed in the 2013 Pb ISA, Pb uptake into human RBCs occurs through a passive
anion transport mechanism, and once Pb is in the cell, little leaves (Bcrgdahl et al.. 1997; Simons. 1993;
Simons. 1986). While the precise mechanisms responsible for decreasing RBC lifespan and mobility are
unknown, occupational Pb exposure has been shown to decrease intracellular free Ca+2 levels and
decrease Ca2+/Mg2+ ATPase activity in RBCs in workers (Abam et al.. 2008; Quintanar-Escorza et al..
2007). These changes are associated with fragility and morphological alterations in RBCs in Pb-exposed
workers. Pb-induced increases in intracellular Ca2+ levels also play a role in the activity of phospholipid
scramblases and flippases in RBCs, increasing access to PS by tissue macrophages and triggering splenic
sequestration and destruction of RBCs, leading to reduced numbers of RBCs in circulation (Jang et al..
2011). Importantly, Ahvavauch et al. (2018) showed that inhibiting Ca2+ increase stimulated by Pb results
in decreased flippase activity and prevented destruction of RBCs. This pathway is depicted in Figure 7-1
by solid lines linking increased intracellular Ca2+ to increased PS exposure and decreased RBC survival.
In addition, phagocytosis of Pb-exposed RBC by human renal proximal tubular cells was mediated by PS
(Kwon and Chung. 2016). Heme-regulated eIF2a kinase was shown to protect RBC from Pb-induced
hemolytic stress in mice (Wang et al.. 2015). Pb-induced hemolysis was also documented in other recent
studies (Hossain et al.. 2015; Mrugesh et al.. 2011). Furthermore, Pb exposure reduced the number of
RBCs and Hb levels in rats (Ibrahim et al.. 2012). Consistent with the pattern seen in epidemiology
studies, the effects of Pb exposure on RBC number and Hb level were more pronounced in 3-month-old
Wistar rats than in adult animals (Daku et al.. 2019). In addition, Cai et al. (2018) reported that Pb
administration reduced RBC lifespan in mice. These findings support the conclusion that Pb alters RBC
survival and function, consistent with the larger body of evidence showing measures of decreased
hematological parameters (i.e., RBC number, Hb, Hct, MCV, and/or MCH) in children (Guo et al.. 2021;
Kuang et al.. 2020; Li et al.. 2018; Liu et al.. 2015; Liu et al.. 2012) and animals (Odo et al.. 2020;
Andielkovic et al.. 2019; Cai et al.. 2018; Corsetti et al.. 2017; Lakshmi et al.. 2013; Berrahal et al.. 2011;
Sharma et al.. 2010; Wang et al.. 2010b; Baranowska-Bosiacka et al.. 2009; Masso-Gonzalez and
Antonio-Garcia. 2009; Simsek et al.. 2009; Masso et al.. 2007).
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Pb exposure also has the potential to disrupt normal hematopoiesis. Erythropoietin (EPO) is a
glycoprotein hormone excreted by the kidney to promote the development of RBCs in bone marrow. As
reviewed in the 2006 Pb AQCD, Pb exposure has been observed to alter EPO production in children
(Graziano et al., 2004; Factor-Litvak et al., 1999; Factor-Litvak et al., 1998). Available data support the
postulation that observed increases in EPO in younger children reflect bone marrow hyperactivity to
counteract RBC destruction, whereas the lack of EPO elevation in older children may reflect a transitional
period in which increasing renal and bone marrow toxicity leads to observed decreases in EPO later in
life (U.S. EPA, 2006). In addition to altering levels of a key hormone involved in hematopoiesis, Pb
exposure has the potential to alter hematopoiesis by causing cytotoxicity (Alghazal et al„ 2008; Marques
et al., 2006; Celik et al., 2005) and senescence (Cai et al., 2018; Nagano et al., 2015) of RBC precursors.
Baktybaeva (2011) reported that intraperitoneal injection of Pb acetate resulted in reduced bone marrow
hematopoiesis in mice. Furthermore, Pb is known to reduce erythropoiesis, causing anemia in children
(Dai et al„ 2017; Ahamed et al., 2011). Altered EPO levels and RBC precursor cytotoxicity have the
potential to alter the number of RBCs in circulation which may lead toto anemia.
7.5.2. Altered Heme Synthesis
Although the mechanisms that could lead to Pb-induced anemia are not fully understood, as
reviewed in the United States Environmental Protection Agency's (U.S. EPA's) 2006 Pb AQCD (U.S.
EPA, 2006), Pb is known to act directly on two enzymes involved in heme synthesis: ALAD and
ferrochelatase. ALAD, a cytoplasmic enzyme requiring zinc (Zn) for enzymatic activity, catalyzes the
rate-limiting step in heme biosynthesis. Inhibition of ALAD activity has been reported in adults (Wang et
al., 2010a) and children (Dai et al., 2017; Wang et al., 2010a; Ahamed et al., 2007) as well as in animal
toxicology studies reporting BLLs relevant to this ISA (i.e., 24.7 (ig/dL) (Rendon-Ramirez et al„ 2007;
Terayama et al., 1986). ALAD activity was also reduced in animal studies reporting BLLs higher than
those meeting the PECOS criteria in this ISA (Mani et al., 2020; Velaga et al., 2014; Whittaker et al.,
2011; Gautam and Flora, 2010; Lee et al., 2005). Pb exposure has also been shown to inhibit
ferrochelatase, a mitochondrial iron (Fe)-sulfur (S)containing enzyme that incorporates Fe2+into
protoporphyrin IX to create heme (Rendon-Ramirez et al., 2007). Pb inhibits the insertion of Fe2+into the
protoporphyrin ring and instead, Zn is inserted into the ring creating Zn-protoporphyrin (ZPP). Evidence
for altered heme synthesis is supported by a large body of evidence collected from occupationally
exposed adults (Ukaejiofo et al., 2009; Khan et al„ 2008; Patil et al., 2006; Karita et al., 2005), children
(Queirolo et al., 2010; Shah et al., 2010; Olivero-Verbel et al., 2007; Riddell et al., 2007), and Pb-exposed
experimental animal models (Andjelkovic et al„ 2019; Cai et al„ 2018; Berrahal et al., 2011;
Baranowska-Bosiacka et al., 2009; Masso-Gonzalez and Antonio-Garcia, 2009; Simsek et al., 2009;
Rendon-Ramirez et al., 2007; Marques et al., 2006; Terayama et al„ 1986) reporting decreased Hb
concentrations in association with Pb exposure or increased BLLs. Further demonstrating the role Pb
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plays in the activity of these important enzymes, chelation therapy restored ALAD activity and reduced
ZPP levels in blood harvested from rats exposed to Pb via drinking water for 90 days (Ata et al.. 2018).
Oxidative stress is caused by an imbalance between production and elimination of reactive
oxygen species (ROS) in cells or tissues that exceed the capacity of antioxidant defense mechanisms.
ROS are unstable, highly reactive molecules formed from molecular oxygen and include, for example,
superoxide (O2), hydroxyl radical (OH), and hydrogen peroxide (H2O2). Although ROS play an important
role in healthy biological systems (e.g., cell signaling, cellular differentiation, immune responses),
unregulated ROS can cause direct and indirect damage to nucleic acids, proteins, and lipids leading to
cytotoxicity, tissue injury, and even disruption of normal physiology (Auten and Davis. 2009). Oxidative
stress is involved in both arms of the pathway leading to anemia shown in Figure 7-1, including effects on
RBC MDA, SOD, CAT, GSH, and GPx levels (Kharoubi et al.. 2008; Lee et al.. 2005). Effects of Pb
exposure on measures of oxidative stress—including MDA and decreased SOD, CAT, GR, and GSSG—
were also observed in vitro (Ciubar et al.. 2007; Coban et al.. 2007).
Supporting the role of oxidative stress in the development of anemia, administration of
antioxidants reduced the effects of Pb exposure on levels of GSH (Alcaraz-Contreras et al.. 2011). MDA
(Alcaraz-Contreras et al.. 2011) and Hb (Farooq et al.. 2016; Saiitha et al.. 2016; Sarkar et al.. 2015;
Eshginia and Marjani. 2013) and reduced the effects of Pb exposure on SOD activity (Eshginia and
Mariani. 2013). ROS production (Nagano et al.. 2015; Sarkar et al.. 2015). hematopoietic stem cell (HSC)
number (Nagano et al.. 2015). HSC colony formation (Cai et al.. 2018). HSC senescence markers (Cai et
al.. 2018; Nagano et al.. 2015). RBC number (Farooq et al.. 2016; Saiitha et al.. 2016; Sarkar et al.. 2015).
and PS exposure (Sarkar et al.. 2015) in animal studies. Conflicting evidence on the effects of antioxidant
treatment on ALAD activity was reported in the literature (Saiitha et al.. 2016; Alcaraz-Contreras et al..
2011). whereas the evidence for effects of Pb exposure on Hb levels was consistent; thus, a solid arrow
connects oxidative stress to decreased RBC Hb content in Figure 7-1. Further demonstrating the role of
oxidative stress in the development of anemia, decreased ALAD activity results in the accumulation of
5-aminolevulinic acid (5-ALA) in blood and urine, where it undergoes tautomerization and autoxidation.
Oxidized 5-ALA leads to the generation of ROS (i.e., O2, OH, H2O2, and an aminolevulinic acid [ALA])
radicals (Hermes-Lima et al.. 1991; Monteiro et al.. 1991; Monteiro et al.. 1989; Monteiro et al.. 1986).
Reflecting the strength of the evidence described above, the pathway connecting oxidative stress to
altered hematopoiesis, cytotoxicity/senescence RBC precursor cells, and decreased RBC survival is
depicted as a solid line.
Decreased RBC Hb content and oxidative stress associated with Pb exposure have been
demonstrated to alter RBC function. This conclusion is supported by direct evidence for binding of Pb to
key enzymes in the heme synthesis pathway, Pb-induced oxidative stress resulting in decreased RBC Hb
content and effects on hematopoiesis, RBC precursor cells, and RBC survival. Decreased RBC number
coupled with impaired RBC function, if of sufficient magnitude, leads to Pb-induced anemia.
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7.6
Summary and Causality Determination
7.6.1. Causality Determination for Red Blood Cell Survival and Function
The 2013 Pb ISA presented causality determinations for two groups of hematological endpoints:
heme synthesis and RBC survival and function. Although there are enzymes and hematological
parameters that are distinct indicators of these processes, the potential biological plausibility pathways in
which exposure to Pb may result in hematological effects demonstrate a spectrum of events, which can be
challenging to attribute to a unique line of evidence (Figure 7-1). For example, altered heme synthesis can
decrease Hb levels, which in turn has been shown to alter RBC function. Because of this
interconnectedness, this assessment presents a single causality determination for Pb exposure and heme
synthesis and RBC survival and function. This approach allows for a more holistic evaluation of inter-
related health endpoints, including a discussion of how all individual lines of evidence contribute to the
overall causality determination. The key evidence, as it relates to the causal framework, is outlined below,
and summarized in Table 7-1.
7.6.2. Evidence for Red Blood Cell Survival and Function
The 2013 Pb ISA concluded that there is a "causal relationship" between Pb exposure and
decreased RBC survival and function (U.S. EPA. 2013). This causality determination was made on the
basis of a strong body of evidence from experimental animal studies demonstrating that Pb exposures
alter several hematological parameters (e.g., Hb, Hct, MCV, MCH), induce oxidative stress (e.g., alter
antioxidant enzyme activities [SOD, CAT, GPx], decrease cellular GSH, and increase lipid peroxidation),
and increase cytotoxicity in RBC precursor cells in rodents exposed to various forms of Pb via drinking
water and gavage resulting in BLLs <30 (ig/dL (Molina et al.. 2011; Baranowska-Bosiacka et al.. 2009;
Lee et al.. 2005). Consistent results were observed in several additional studies in rodents that did not
report BLLs. Epidemiologic evidence was coherent with results from the evaluated toxicological studies
but was subject to more uncertainties. Notably, the epidemiologic evidence consisted of cross-sectional
studies that were conducted in populations with higher mean Pb exposures (i.e., BLLs >10 |ig/dL). did
not thoroughly consider potential confounders, and lacked rigorous statistical methodology. These
limitations precluded strong conclusions on the directionality of effects; the level, timing, frequency, and
duration of Pb exposure that contributed to the observed associations; and whether the observed
associations are independent of potential confounders. Although there were substantial uncertainties in
the epidemiologic evidence, animal toxicological evidence established a clear basis for temporality of
exposure to Pb and effects on RBCs. Additional support for these findings was provided by toxicological
and epidemiologic studies demonstrating increased intracellular calcium concentrations, decreased
Ca2+/Mg2+ ATPase activity, and increased PS exposure, establishing biologically plausibility for Pb-
induced changes in RBC survival.
7-17
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Although limited in number, recent PECOS-relevant animal toxicological studies continue to
support the findings from the last review. The most consistent evidence comes from studies that report
decreased Hb levels in rodents following Pb exposures, resulting in BLLs ranging from 7.5 to 14.7 (ig/dL
(Andielkovic et al.. 2019; Cai et al.. 2018; Berrahal et al.. 2011). Other recent toxicological studies noted
Pb-induced decrements in Hct (Andielkovic et al.. 2019). packed cell volume (PCV) (Berrahal et al..
2011). and hematopoiesis (Andielkovic et al.. 2019). Recent epidemiologic studies expand on the
evidence presented in the 2013 Pb ISA and are coherent with the experimental evidence. Although the
recent studies are also cross-sectional, they include populations with much lower BLL means (<10 (ig/dL)
and include more robust adjustment for potential confounding, addressing important uncertainties from
the last review. The most consistent epidemiologic evidence indicates an inverse association between
BLLs and Hb levels in children (Section 7.3.1), which is in line with the evidence from recent
experimental animal studies. While the clinical relevance of small mean decrements in Hb across
exposure quintiles is unclear, a few of the recent epidemiologic studies observed higher odds of prevalent
anemia in children associated with higher quantiles of BLLs (Guo et al.. 2021; Li et al.. 2018). Recent
epidemiologic studies of Hb in adults were more limited in number and less consistent than those in
children. Additionally, the relatively few studies examining RBC counts were also inconsistent.
7.6.3. Evidence for Heme Synthesis
The 2013 Pb ISA concluded there is a "causal relationship" between Pb exposure and altered
heme synthesis (U.S. EPA, 2013). This determination was based on a small but consistent body of studies
in adult animals reporting that Pb exposures via drinking water and gavage (resulting in BLLs relevant to
this ISA) for 15 days to 9 months decreased ALAD (Rendon-Ramirez et al.. 2007; Terayama et al.. 1986)
and ferrochelatase activities (Rendon-Ramirez et al.. 2007). Notably, Rendon-Ramirez et al. (2007)
observed effects on ALAD and ferrochelatase activities in albino Wistar rats at mean BLLs of 24.7 (ig/dL
after Pb administration drinking water for 15 or 30 days. Supporting this toxicological evidence was a
larger body of ecotoxicological studies that demonstrate altered heme synthesis in Pb-exposed aquatic and
terrestrial invertebrates and vertebrates. Ecological evidence from previous reviews consistently observed
Pb-induced ALAD inhibition in multiple species, including birds and fish (U.S. EPA, 2013, 2006). Cross-
sectional epidemiologic studies provided supporting evidence that concurrent elevated BLLs are
associated with lower ALAD and ferrochelatase activities in both adults and children. However, the
majority of these studies are limited by the lack of rigorous methodology and consideration of potential
confounding. Although there were limitations in the epidemiologic evidence, some studies in children did
control for or consider potential confounding, and effects in adults and children in these studies are
coherent with effects observed in animal toxicological studies.
The relationship between Pb exposure and altered heme synthesis was further supported by cross-
sectional epidemiologic studies indicating an inverse association between BLLs and Hb in children and
occupationally exposed adults. These findings were consistent with several toxicological studies that
7-18
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observed decreased Hb levels in laboratory animals exposed to Pb. Decreased Hb levels can be a direct
indicator of decreased heme synthesis.
Recent PECOS-relevant studies are limited in number and focus mainly on Hb levels but continue
to provide support for Pb-related alterations in heme synthesis. Notably, recent epidemiologic studies
indicating an association between higher BLLs and lower Hb include more robust statistical methods,
expanded consideration of potential confounders, and populations with much lower BLLs than the studies
included in the previous reviews (mean or median BLLs ranging from 3.04 to 8.38 (ig/dL; Section 7.3.1).
The recent epidemiologic evidence is coherent with recent toxicological studies, which observed Hb
decrements in Pb-exposed mice (Andielkovic et al.. 2019; Cai et al.. 2018). While the cross-sectional
nature of the epidemiologic studies introduces uncertainty about the temporality of the exposure and
outcome, animal toxicological evidence establishes clear temporality of exposure to Pb and altered heme
synthesis.
7.6.4. Causality Determination
In summary, there is sufficient evidence to conclude that there is a causal relationship
between Pb exposure and hematological effects, including altered heme synthesis and decreased
RBC survival and function. The strongest support for this causality determination comes from
experimental animal studies demonstrating that exposures to various forms of Pb via drinking water or
gavage resulting in BLLs < 30 (ig/dL, alter several hematological parameters (e.g., Hb, Hct, MCV, MCH)
and decrease ALAD and ferrochelatase activities in rodents. These toxicology results are coherent with
findings from epidemiologic studies that report Pb exposures alter key hematological parameters and
enzymes. Although all evaluated epidemiologic studies are cross-sectional, toxicological studies establish
temporality between exposure to Pb and effects on heme synthesis and RBCs. Additionally, recent
epidemiologic studies address uncertainties from previous reviews by expanding adjustment for potential
confounders and using more robust statistical methods (i.e., multivariable regression models). Because of
the contribution of bone Pb levels to concurrent BLLs, associations with concurrent BLLs may reflect an
effect of past and/or recent Pb exposures. Therefore, there is uncertainty regarding the timing, duration,
and level of Pb exposure associated with observed hematological effects in children and adults. Biological
plausibility for the observed associations is provided by toxicological and epidemiologic studies
demonstrating increased intracellular calcium concentrations, decreased Ca2+/Mg2+ ATPase activity, and
increased PS exposure, which collectively can lead to fragility, morphological alterations in RBCs, and
RBC destruction.
7-19
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Table 7-1 Summary of evidence indicating a causal relationship between Pb exposure and hematological
effects
Rationale for Causality
Determination3
Key Evidence"
Key References"
Pb Biomarker Levels Associated
with Effects0
Red Blood Cell Survival and Function
Large body of studies with generally
consistent findings for decreased RBC
survival and function in rodents:
See Section 7.3.2
Mean BLLs (±SD):
Decreased plasma Hb concentration
Cai etal. (2018)
9.3 ± 0.98 |jg/dL
Consistent evidence from
toxicological studies with
relevant exposures
Andielkovic et al. (2019)
Berrahal etal. (2011)
Sharma et al. (2010)
Masso-Gonzalez and Antonio-Garcia
(2009)
Baranowska-Bosiacka et al. (2009)
29.0 ± 43.1 |jg/dL
12.67 ±1.68 |jg/dL PND 40;
7.49 ± 0.78 |jg/dL PND 65
1.72 ± 0.02 |jg/dL
22.8 ± 0.50 |jg/dL
7.11 ± 1.7 |jg/dL
Decreased Hct
Andielkovic et al. (2019)
Masso et al. (2007)
29.0 ± 14.7 |jg/dL
22.8 ± 0.50 |jg/dL
7-20
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Rationale for Causality
Determination3
Key Evidence"
Key References"
Pb Biomarker Levels Associated
with Effects0
Decreased RBCs
Decreased PCV
Increased eryptosis
Decreased hematopoiesis
Increased oxidative stress
Masso-Gonzalez and Antonio-Garcia
(2009)
Berrahal etal. (2011)
Andielkovic et al. (2019)
Cai etal. (2018)
Sharma et al. (2010)
U.S. EPA (2013)
U.S. EPA (2013)
U.S. EPA (2006)
U.S. EPA (2013)
U.S. EPA (2006)
In vitro:
Ciubar et al. (2007)
Coban et al. (2007)
22.8 ± 0.50 |jg/dL
12.67 ±1.68 |jg/dL PND 40;
7.49 ± 0.78 |jg/dL PND 65
29.0 ± 14.7 |jg/dL
9.3 ± 0.98 |jg/dL
1.72 ± 0.02 |jg/dL
7-21
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Rationale for Causality
Determination3
Key Evidence"
Key References"
Pb Biomarker Levels Associated
with Effects0
Shin etal. (2007)
Consistent evidence from
multiple epidemiologic
studies of children with
relevant BLLs provides
coherence with
toxicological evidence.
Cross-sectional studies provide support for
experimental evidence with consistent
associations between blood Pb and
decreased in Hb in children. Recent studies
adjusted for a number of relevant potential
confounders, including age, sex, BMI,
factors, and nutrition.
Guo et al. (2021)
Kuanq et al. (2020)
Li etal. (2018)
Liu etal. (2015)
Liu etal. (2012)
Mean BLLs:
3.07-3.21 |jg/dL
3.04 |jg/dL
8.38 |jg/dL (Median)
7.33 |jg/dL
4.30 |jg/dL (Median)
Cross-sectional studies provide generally U.S. EPA (2013)
consistent evidence for associations between
blood Pb and altered hematological
parameters (e.g., RBC counts, Hct, MCV, and
MCH), measures of oxidative stress
(e.g., SOD, CAT, GPx, GSH, and lipid
peroxidation), and hematopoiesis
(e.g., decreased erythropoietin). Evidence
base limited by lack of adjustment for potential
confounders and populations with higher
BLLs.
Biological Plausibility
Altered RBC membrane ion Evidence of increased [Ca2+]i and decreased See Section 7.5.2
transport Ca2+/Mg2+ATPase activity in the RBCs of
exposed workers. [Ca2+]i levels highly
correlated with blood Pb even among
unexposed controls.
7-22
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Key References"
Pb Biomarker Levels Associated
with Effects0
Rationale for Causality
Determination3
Key Evidence"
Altered RBC membrane ion
transport
[Ca2+]i levels increased in RBCs from healthy
volunteers when exposed in vitro to Pb.
[Ca2+]i associated with increased RBC fragility
and alterations in RBC morphology.
PS exposure Consistent evidence from in vivo and in vitro
studies that Pb exposure increases PS
exposure on RBC membranes via modulation
of[Ca2+]i concentrations. Increased PS
exposure leads to eryptosis and phagocytosis
by macrophages.
Heme Synthesis
Consistent evidence in
animals with relevant
exposures
A small, but consistent toxicology evidence
base indicates decreased heme synthesis in
rodents with relevant Pb concentrations and
routes of exposure.
Rendon-Ramirez et al. (2007)
BLL: 24.7 ± 2.4 pg/dL
Exposures: 500-5,000 ppm in drinking
water, 15-30 days as adults
Teravama et al. (1986)
Coherence in a limited
number of epidemiologic
studies with relevant BLLs
Cross-sectional studies that considered
potential confounding by age, sex, urban/rural
residence, height, weight, BMI found
consistent associations with lower ALAD and
ferrochelatase activities in children.
Ahamed et al. (2006)
Ahamed et al. (2007)
Mean BLL: 7.40 and 13.27 pg/dL
BLL: >10 pg/dL compared with
<10 pg/dL
Concurrent BLL associated with lower ALAD Wang et al. (2010a)
and higher ZPP in adults with consideration
for potential confounding by age, sex, smoking
status, and alcohol use.
Mean BLL: 6.71 pg/dL
7-23
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Rationale for Causality
Determination3
Key Evidence"
Key References"
Pb Biomarker Levels Associated
with Effects0
Support from toxicological Consistent evidence in animals with relevant Baranowska-Bosiacka et al. (2009).
and epidemiologic
evidence for decreases in
Hb, a direct marker of
decreased heme synthesis
Pb exposures for decreases in Hb levels.
Sharma et al. (2010)
Section 7.4.2
Consistent associations between concurrent
BLLs and decreased Hb in children.
Associations observed at low BLLs with
thorough consideration of potential
confounders.
Guo et al. (2021)
Kuana et al. (2020)
Li et al. (2018)
Liu et al. (2015)
Liu et al. (2012)
Adult animals: BLL7.11 ± 1.7 pg/dL
after 9-mo Pb exposure
Adult animals: BLL 1.7 pg/dL after 40-
day Pb exposure
Mean BLLs:
3.07-3.21 pg/dL
3.04 pg/dL
8.38 pg/dL (Median)
7.33 pg/dL
4.30 pg/dL (Median)
Biological Plausibility Altered Ion Status: Evidence that Pb See Section 7.5.2
competitively inhibits the binding ofZn ions
necessary for ALAD activity. Pb also inhibits
the incorporation of Fe2+ into protoporphyrin IX
by ferrochelatase, resulting in Zn-
protoporphyrin production.
ALAD = 6-aminolevulinic acid dehydratase; BLL = blood lead level; BMI = body mass index; Ca2+ = calcium ion(s); CAT = catalase; Fe2+ = iron; GPx = glutathione peroxidase;
GSH = glutathione; Hb = hemoglobin; Hct = hematocrit; i = inorganic; MCV = mean corpuscular volume; Mg2+ = magnesium; Pb = lead; PCV = packed cell volume; PND = postnatal
day; PS = phosphatidylserine; RBC = red blood cell; SES = socioeconomic status; SOD = superoxide dismutase; Zn = zinc; ZPP = Zn-protoporphyrin.
"Based on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
'Describes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
Describes the Pb biomarker levels at which the evidence is substantiated.
7-24
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7.7 Evidence Inventories—Data Tables to Summarize Study Details
Table 7-2 Epidemiologic studies of exposure to Pb and hematological effects
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Children
tGuoetal. (2021)
Guangdong Women and
Children's Hospital
Blood
Hb
Age and sex
Hb
Mean Difference (g/L)
Guangdong
China
2014-2017
Cross-sectional
n: 17,486
Children 0-5 yr old visiting
hospital for routine health
examination
BLLs were measured using
atomic absorption
spectrometry
Age at Measurement:
0-5 yr
Hb (g/L) measured
using an automated
hematology analyzer
Age at Outcome:
0-5 yr
Q1 (<1.61)
Reference
Q2 (1.61-2.44)
-0.05 (-0.51, 0.40)
Q3 (2.44-3.33)
-0.02 (-0.48, 0.43
Means:
Males: 3.21 |jg/dL;
Females: 3.07 |jg/dL
Q4 (3.33-4.50)
-0.48 (-0.94, -0.04)
Q5(>4.50)
-1.25 (-1.71, -0.78)
Anemia (OR)
Q1 (<1.61)
Reference
Q2 (1.61-2.44)
1.08 (0.94, 1.23)
Q3 (2.44-3.33)
1.16 (1, 1.33)
Q4 (3.33-4.50)
1.25 (1.07, 1.43)
Q5(>4.50)
7-25
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
1.45 (1.26, 1.67)
tKuana et al. (2020) n: 395
Nanjing
China
2012
Cross-sectional
Convenience sample of
children 7-11 yr old
Blood
Blood Pb was measured in
venous whole blood using
ICP-MS
Age at Measurement:
7-11 yr old
Mean: 3.04 |jg/dL;
Median: 2.61 |jg/dL
Hematological
Parameters
RBC, Hb, Hct, MCV,
MCH, and MCHC
measured by a whole
cell analyzer
Age at Outcome:
7-11 yr old
Picky eaters and
passive smoking (age,
gender, parents'
education, and parents'
occupation also
considered)
RBC Count (1012/L)
Boys: 0.02 (-0.01, 0.04)
Girls: 0.01 (-0.02, 0.03)
Hb (g/L)
Boys:-0.12 (-0.22,
-0.02)
Girls: -0.08 (-0.23, 0.07)
Hct (%)
Boys: -0.04 (-0.08,
-0.01)
Girls: -0.02 (-0.07, 0.02)
MCV (fL)
-0.04 (-0.06, -0.02)
MCH (pg)
-0.01 (-0.02, -0.00)
MCHC (g/L)
0.26 (-0.64, 1.15)
tLiuetal. (2012)
China Jintan Child Cohort
Study
Blood
Hb
Age, sex, height,
weight, iron deficiency
Hb (g/dl_)*
Full Population
Changzhou City
n: 140
Blood Pb was measured in
Hb measured in whole
-0.096 (-0.18, -0.012)
China
whole blood using GFAAS
blood using a
Blood Pb < 10 ijg/dL
Cross-sectional
Convenience sample of
preschool age children
Age at Measurement:
Median age: 3 yr old
Median: 4.3 |jg/dL
Maximum: 11.4 pg/dL
photoelectric
colorimeter
Age at Outcome:
Median age: 3 yr old
-0.174 (-0.27, -0.078)
7-26
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tLi etal. (2018)
Hubei and Hunan
Provinces
China
2012-2017
Cross-sectional
Blood Lead Intervention
Program
n: 758
Children Ages 5-8 yr
recruited from four counties
in two provinces
One county in each province
had high environmental Pb
levels (battery plant and
mining)
Blood
Blood Pb was measured in
venous whole blood using
GFAAS
Age at Measurement:
5-8 yr
Median: 8.38 |jg/dL
75th: 13.51 pg/dL
90th: 18.77 pg/dL
95th: 21.82 pg/dL
Hematological
Parameters
Hb, MCH, RBCs, and
Pits measured in
venous whole blood
using an automated
hematology analyzer
Age at Outcome:
5-8 yr
Age, sex, BMI,
environmental Pb
exposure level, and
serum iron, zinc, and
calcium
ORs
Decreased Hb (<115 g/L)
1.05 (1.00, 1.11)
Decreased RBC
(<4 x 1012/L for boys;
<3.5 x 1012/L for girls)
1.11 (1.05, 1.16)
Decreased Pit
(<100 x 109/L)
1.11 (1.05, 1.16)
Decreased MCH (<27 pg)
1.11 (1.05, 1.16)
tLiuetal. (2015)
n: 855
Blood
Hb Age, sex, residence
area, and SES
Hb (g/L)
Mean Difference
Guiyu, Chendian,
Children 3-7 yr old from e-
Blood Pb was measured
Hb, MCH, RBCs, and
Q1 (5.98-13.52)*
and Chaonan
waste processing area or
using GFAAS. Blood Pbwas
Pits measured in
Reference
China
control industrial areas
divided by Hct as a fraction of
venous whole blood
Q2 (13.52-19.35)*
2006-2011
without high environmental
the whole blood to estimate
using an automated
Pb exposures
erythrocyte Pb
hematology analyzer
-0.02 (-1.89, 1.52)
Cross-sectional
Q3 (19.35-28.42)*
Age at Measurement:
Age at Outcome:
-3.01 (-4.71, 1.31)
3-7 yr old
3-7 yr old
Q4 (28.42-101.01)*
-3.97 (-5.68, -2.27)
Median:
Blood Pb: 7.33 pg/dL;
*Erythrocyte Pb (pg/dL)
Erythrocyte Pb: 19.3 pg/dL
Adults
7-27
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tPark and Lee
(2013)
South Korea
2008-2010
Cross-sectional
KNHANES
n: 4522
General population, >20 yr
old
Blood
Blood Pb was measured in
venous whole blood using
GFAAS
Age at Measurement:
>20 yr
Geometric Means:
Males: 2.46 |jg/dL;
Females: 1.98 |jg/dL
Hb
Blood Hb (g/dL)
measured using an
automated hematology
analyzer
Age at Outcome:
>20 yr
Age, BMI, education, Hb (g/dL)*
smoking and drinking
status, and rural/urban
residence
Men
0.04 (0.03, 0.06)
Women
0.04 (0.02, 0.06)
*Not standardized.
In(Pb) increase
Per
tKim and Lee
(2013)
South Korea
2008-2010
Cross-sectional
KNHANES
n: 5951
General population, >20 yr
old
Blood
Blood Pb was measured in
whole blood using GFAAS.
Blood Pb was divided by Hct
as a fraction of the whole
blood to estimate erythrocyte
Pb
Age at Measurement:
>20 yr old
Median:
Blood Pb: 2.31 pg/dL;
Erythrocyte Pb: 5.4 pg/dL
75th:
Blood Pb: 3.01 pg/dL;
Erythrocyte Pb: 6.9 pg/dL
Hb
Hb measured in whole
blood using an
automated hematology
analyzer
Age at Outcome:
>20 yr old
Sex, age, obesity,
residence area,
education level,
smoking and drinking
status, serum ferritin,
and serum creatinine
Hb (g/L)
Mean Difference
Q1 (<1.73)
Reference
Q2 (1.73-2.31)
0.13 (0.03, 0.23)
Q3 (2.31-3.01)
0.33 (0.23, 0.42)
Q4 (>3.01)
0.42 (0.30, 0.53)
Q1 (<4.1)*
Reference
Q2 (4.1-5.4)
-0.06 (0.15, 0.03)
Q3 (5.4-6.9)
-0.06 (-0.15, 0.03)
Q4 (>6.9)
-0.14 (-0.25, -0.04)
7-28
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
*Erythrocyte Pb (pg/dL)
tLa-Llave-Leon et n: 292
al. (2015)
Durango
Mexico
2007-2008
Cross-Sectional
Pregnant women, 14-41 yr
old
Blood
Blood Pb was measured in
venous whole blood using
GFAAS
Age at Measurement:
14-41 yr old
Hematological
Parameter
RBC, Hb, Hct, MCV,
MCH, and MCHC
measured using an
automated hematology
analyzer
BMI, gestational age,
age, parity, gestations,
and household monthly
income per person
RBC Count (x106Mg/dL)
0.034 (0.013, 0.056)
Mean: 2.79 |jg/dL
Age at Outcome:
14-41 yr old
BMI = body mass index; BW = body weight; CI = confidence interval; e-waste = electronic waste; GFAAS = graphite furnace atomic absorption spectrometry; Hb = hemoglobin;
Hct = hematocrit; ICP-MS = inductively coupled plasma mass spectrometry; KNHANES = Korea National Health and Nutrition Examination Survey; In = natural log; MCH = mean
corpuscular hemoglobin; MCHC = mean corpuscular hemoglobin concentration; MCV = mean corpuscular volume; OR = odds ratio; Pb = lead; Pit = platelet; PND = postnatal day;
Q = quartile; RBC = red blood cell; RDW = red blood cell distribution width; SES = socioeconomic status; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb level or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results
corresponding to a change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized
accordingly.
fStudies published since the 2013 Pb ISA.
7-29
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Table 7-3 Animal toxicological studies of Pb exposure and hematological effects
Study (StolEin), n, Timing of Exposure Details BLL as Reported (pg/dL)3
' v Sex ' Exposure (Concentration, Duration) ^ VMa ' Examined
Berrahal et Rat (Wistar)
al. (2011)
Control (vehicle),
M,n = 12-16
50 mg/L Pb, M,
n = 12-16
PND 1 to
PND21:
Lactational
PND 21 to
PND 40 or
PND 65:
Drinking water
Dams were given 50 mg/L Pb acetate in drinking PND 40:
Hct, Hb
water until weaning on PND 21. Male offspring
received 50 mg/L Pb acetate in drinking water
from PND 21 to PND 40 or PND 65. Control
animals received tap water.
1.76 ± 0.33 |jg/dL for 0 pg/dL
12.67 ± 1.68 pg/dL for 50 mg/L
PND 65:
2.06 ± 0.35 pg/dL for 0 pg/dL
7.49 ± 0.78 pg/dL for 50 mg/L
Basha et al. Rat (Wistar) PND 1
(2012) Control (vehicle), *° P^D 21
M, n = 8
0.2% Pb, M, n = 8
PND 12 mo:
0.56 ± 0.08 pg/dL for 0%,
16.4 ± 1.95 pg/dL for 0.2%,
7.2 ± 0.56 pg/dL for 0.2% +
supplementation
Damns given Pb acetate in drinking water or Pb
acetate containing water supplemented with
0.02% calcium, zinc, and iron. Control group
received deionized water as vehicle (no
supplement).
Pups exposed through lactation.
PND 45: RBC, Hb
0.42 ± 0.04 pg/dL for 0%,
52.5 ± 0.67 pg/dL for 0.2%,
21.1 ± 1.12 pg/dL for 0.2% +
supplementation
PND 24 mo:
0.46 ± 0.02 pg/dL for 0%,
12.2 ± 0.76 pg/dL for 0.2%,
4.8 ± 0.5 pg/dL for 0.2% for
0.2% + supplementation
Zou et al. Mouse (ICR) 3 wk exposure Rats received 250 mg/L Pb acetate in redistilled PND 58: RBC, MCHC
(2015) Control (vehicle), drinking water for 3 wk. The rats were 30 d old 1.8 pg/dL for 0 mg/L
m n-m when acquired, but the authors did not specify
M'n"1° the age at the time of treatment. 21.7 pg/dL for 250 mg/L
250 mg/L Pb, M,
n = 10
7-30
-------
Species
Timing of
Study (Stock/SJrain), n, ¦ —»-
Exposure Details
(Concentration, Duration)
BLL as Reported (|jg/dL)a
Endpoints
Examined
Corsetti et al. Mouse
(2017) (C57BL.6)
Control (vehicle),
M, n = 8
d 30 to d 75 Mice were exposed via drinking water for 45
consecutive days. Control animals were exposed
to drinking water containing acetic acid (1 mL/L).
<5 |jg/dL for 0 ppm
21.6 |jg/dL for 200 ppm
RBC, Hb, Hct,
MCV, MCH,
MCHC, RDW %,
Pits
200 ppm Pb, M,
n = 8
Andielkovic Rat (Wistar) NR
et al. (2019) control (vehicle),
M, n = 8
0.2% Pb, M, n = 6
Rats (250 g), age at time of dosing not reported,
were exposed to a single dose of 150 mg Pb/kg
BW Pb acetate via oral gavage. Control animals
were given "water."
24.9 ± 1 9 |jg/kg for 0 mg Pb/kg BW
(2.6 ± 2.0 |jg/dL)
291.2 ± 139 |jg/kg for 150 mg Pb/kg
BW
(29.0 ± 14.7 |jg/dL)
RBC, Hb, Hct,
MCV, MCH,
MCHC, Pits
Cai et al.
(2018)
Rat (Sprague
Dawley)
Control (vehicle),
M/F, n = 5
0.2% Pb, M/F,
n = 5
8-10 wk to 20- Rats were 8-10 wk old when acquired. Whether
22 wk or not the rats were allowed to acclimate to the
facility prior to study initiation was not reported.
The number of males and females not reported.
Control animals received tap water.
The exposure period was 12 wk, assumed rats
were exposed 7 d/wk for a total of 84 d.
20.5 ± 0.68 |jg/L for 0%
(2.2 ± 6.4 |jg/dL)
87.4 ± 9.2 |jg/L for 0.2%
(9.3 ± 0.98 |jg/dL)
Hb, Pits,
Erythrocyte life
span, RBC
BLL = blood lead level; BW = body weight; d = day; Hb = hemoglobin; Hct = hematocrit; M = male; M/F = male/female; MCH = mean corpuscular hemoglobin; MCHC = mean
corpuscular hemoglobin concentration; MCV = mean corpuscular volume; mo = month(s); NR = not reported; Pb = lead; Pit = platelet; PND = postnatal day; RBC = red blood cell;
RDW = red blood cell distribution width; wk = week(s).
alf applicable, reported values for BLL were converted to |jg/dL using WebPlot Digitizer (https://apps.automeris.io/wpd/) and are shown in parenthesis.
7-31
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dehydratase activity, urinary 5-aminolevulinic acid concentration and zinc protoporphyrin level among
people with low level of lead exposure. Int J Hyg Environ Health 213: 52-58.
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Wang. X: Wang. L: Liu. S. (2015). Heme-regulated eIF2a kinase plays a crucial role in protecting erythroid cells
against Pb-induced hemolytic stress. Chem Res Toxicol 28: 460-469. http://dx.doi.org/10.1021/tx500422q.
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United States
Environmental Protection
Agency
EPA/600/R-23/375
January 2024
www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 8: Reproductive and
Developmental Effects
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
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CONTENTS
DOCUMENT GUIDE 8-iii
LIST OF TABLES 8-v
LIST OF FIGURES 8-vi
ACRONYMS AND ABBREVIATIONS 8-vii
APPENDIX 8 REPRODUCTIVE AND DEVELOPMENTAL EFFECTS 8-1
8.1 Introduction and Summary of the 2013 Integrated Science Assessment 8-1
8.1.1 Effects on Pregnancy and Birth Outcomes 8-2
8.1.2 Effects on Development 8-2
8.1.3 Effects on Female Reproductive Function 8-3
8.1.4 Effects on Male Reproductive Function 8-3
8.2 Scope 8-4
8.3 Effects on Pregnancy and Birth Outcomes 8-5
8.3.1 Maternal Health During Pregnancy 8-6
8.3.2 Prenatal Growth 8-10
8.3.3 Preterm Birth 8-17
8.3.4 Birth Defects 8-20
8.3.5 Spontaneous Abortion and Pregnancy Loss and Fetal and Infant Mortality 8-23
8.3.6 Placental Function 8-25
8.3.7 Other Pregnancy and Birth Outcomes 8-27
8.4 Effects on Development 8-29
8.4.1 Effects on Postnatal Growth 8-29
8.4.2 Effects on Puberty among Females 8-34
8.4.3 Effects on Puberty among Males 8-37
8.4.4 Other Developmental Effects 8-40
8.5 Effects on Female Reproductive Function 8-41
8.5.1 Effects on Hormone Levels and Menstrual/Estrous Cycle 8-41
8.5.2 Effects on Female Fertility 8-45
8.5.3 Effects on Morphology and Histology of Female Sex Organs (Ovaries, Uterus,
Fallopian Tubes/Oviducts, Cervix, Vagina, and Mammary Glands) 8-47
8.6 Effects on Male Reproductive Function 8-48
8.6.1 Effects on Sperm/Semen Production, Quality, and Function 8-48
8.6.2 Effects on Hormone Levels in Males 8-51
8.6.3 Effects on Male Fertility 8-54
8.6.4 Effects on Morphology and Histology of Male Sex Organs 8-55
8.7 Biological Plausibility 8-57
8.7.1 Pubertal Onset 8-58
8.7.2 Male Reproduction Function 8-59
8.8 Summary and Causality Determination 8-61
8.8.1 Summary of Effects on Pregnancy and Birth Outcomes 8-61
8.8.2 Summary of Effects on Development 8-64
8.8.3 Summary of Effects on Female Reproductive Function 8-67
8.8.4 Summary of Effects on Male Reproductive Function 8-70
8.9 Evidence Inventories - Data Tables to Summarize Study Details 8-79
8.10 References 8-243
8-iv
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LIST OF TABLES
Table 8-1 Summary of evidence contributing to causality determinations for Pb exposure and
reproductive and developmental effects 8-74
Table 8-2 Epidemiologic studies of exposure to Pb and maternal health outcomes 8-79
Table 8-3 Animal toxicological studies of Pb exposure and pregnancy and birth outcomes 8-94
Table 8-4 Epidemiologic studies of Pb exposure and prenatal growth 8-99
Table 8-5 Epidemiologic studies of Pb exposure and preterm birth 8-134
Table 8-6 Epidemiologic studies of Pb exposure and birth defects 8-146
Table 8-7 Epidemiologic studies of Pb exposure and fetal and infant mortality and spontaneous
abortion and pregnancy loss 8-153
Table 8-8 Epidemiologic studies of Pb exposure and placental function 8-157
Table 8-9 Epidemiologic studies of Pb exposure and other pregnancy and other birth outcomes 8-159
Table 8-10 Epidemiologic studies of Pb exposure and postnatal growth 8-165
Table 8-11 Animal toxicological studies of Pb exposure and development 8-184
Table 8-12 Epidemiologic studies of exposure to Pb and puberty in females and puberty in males 8-191
Table 8-13 Epidemiologic studies of exposure to Pb and other developmental effects 8-204
Table 8-14 Epidemiologic studies of exposure to Pb and female reproductive effects 8-210
Table 8-15 Animal toxicological studies of Pb exposure and female reproductive effects 8-220
Table 8-16 Epidemiologic studies on exposure to Pb and male reproductive effects 8-223
Table 8-17 Animal toxicological studies of exposure to Pb and male reproductive effects 8-240
8-v
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LIST OF FIGURES
Figure 8-1 Potential biological pathways for reproductive and developmental effects following
exposure to Pb. 8-58
8-vi
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ACRONYMS AND ABBREVIATIONS
2PN oocytes with two pronuclei
AAS atomic absorption spectrometry
AD abdominal diameter
AGD anogenital distance
AGDap anopenile distance
AGDas anoscrotal distance
ALAD S-aminolevulinic acid dehydratase
ALSPAC Avon Longitudinal Study of Parents
and Children
AMH anti-Mullerian hormone
AQCD Air Quality Criteria Document
ART assisted reproductive technology
As arsenic
BKMR Bayesian kernel machine regression
BL birth length
BLL blood lead level
BMI body mass index
BMIZ BMI-for-age Z-score
BT20+ Birth to Twenty Plus
BW birth weight
BWGA birth weight-for-gestational age
BWZ birth weight Z-score
C-ABCS China-Anhui Birth Cohort Study
CANDLE Conditions Affecting Neurocognitive
Development and Learning in Early
Childhood
CC chest circumference
CCG Charlotte-Concord-Gastonia
Cd cadmium
CD cephalic diameter
CHD congenital heart disease
CHECK Children's Health and Environmental
Chemicals in Korea
CHL crown-heel length
CI confidence interval
CMS Charlotte Motor Speedway
Cr chromium
d day(s)
DBP diastolic blood pressure
E2 estradiol
E3G estrone-3-glucuronide
EAAS electrothermal atomic absorption
spectrometry
ELEMENT Early Life Exposure in Mexico to
Environmental Toxicants
ELISA enzyme-linked immunosorbent assay
EMASAR Study on the Environment and
Reproductive Health
e-REACH e-waste Recycling Exposure and
Community Health
ETS environmental tobacco smoke
fE2 free estradiol
FLEHS Flemish Environment and Health Study
FSH follicle stimulating hormone
FT free testosterone
FT3 free triiodothyronine
FT4 free thyroxine
GA gestational age
GD gestational day
GDM gestational diabetes mellitus
GEE generalized estimating equation
GFAAS graphite furnace atomic absorption
spectrometry
GnRH gonadotropin-releasing hormone
GSI Global Severity Index
HAZ height-for-age Z-score
HC head circumference
HCAZ head circumference for age Z-score
hCG human chorionic gonadotropin
HFIAS Household Food Insecurity Access
Scale
Hg mercury
HOME Health Outcomes and Measures of the
Environment
HR hazard ratio
HR-ICP-MS high resolution inductively coupled
plasma mass spectrometry
HTZ height Z-score
hr hour(s)
ICP-AES inductively coupled plasma atomic
emission spectroscopy
ICP-MS inductively coupled plasma mass
spectrometry
ICP-QQQ inductively coupled plasma triple quad
IgE immunoglobulin E
IGF-1 insulin-like growth factor 1
IGT impaired glucose tolerance
IL-33 interleukin-33
INMA Instituto de Nanociencia y Materiales
de Aragon
IQR interquartile range
ISA Integrated Science Assessment
IUGR intrauterine growth restriction
8-vii
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IVF in vitro fertilization
JECS Japan Environment and Children's
Study
K6 Kessler Psychological Distress Scale
KNHANES Korea National Health and Nutrition
Examination Survey
K-XRF K-shell X-ray fluorescence
LA-ICP-MS laser ablation-inductively coupled
plasma-mass spectrometry
LBW low birth weight
LESPW Life Event Scale for Pregnant Women
LGA large for gestational age
LH luteinizing hormone
LIFE Longitudinal Investigation of Fertility
and the Environment
LMP last menstrual period or last missed
period
In natural log
LOD limit of detection
MAL-ED Interactions of Malnutrition and Enteric
Infections: Consequences for Child
Health and Development
Mil metaphase II
min minute(s)
MIREC Maternal-Infant Research on
Environmental Chemicals
miRNA micro RNA
MMP matrix metalloproteinase
Mn manganese
mo month(s)
MOCEH Mothers'and Children's
Environmental Health
MSA Metropolitan Statistical Area
mtDNA mitochondrial DNA
mtDNAcn mitochondrial DNA copy number
NASCAR National Association for Stock Car
Auto Racing
NHANES National Health and Nutrition
Examination Survey
NICE Nutritional impact on Immunological
maturation during Childhood in relation
to the Environment
NR not reported
NS non-stress
NTD neural tube defect
OFC orofacial cleft
OGTT oral glucose tolerance test
OR odds ratio
Pb lead
PECOS Population, Exposure, Comparison,
Outcome, and Study Design
PI Ponderal Index
PIR poverty-income ratio
PM2.5 fine particulate matter
PND postnatal day
PROGRESS Programming Research in Obesity,
Growth, Environment and Social
Stressors
PROM premature rupture of membranes
PROTECT Puerto Rico Test site for Exploring
Contamination Threats
QL lower quartile
QUS quantitative ultrasound
ROS reactive oxygen species
RR relative risk
rTL relative telomere length
SA semen analysis
SBP systolic blood pressure
SCL-90-R Symptom-Checklist-90-Revised
SD standard deviation
Se selenium
SE standard error
SES socioeconomic status
SGA small for gestational age
SHBG sex hormone binding globulin
SNP single nucleotide polymorphism
SPECT Survey on the Prevalence in East China
for Metabolic Diseases and Risk
Factors
T testosterone
T# tertitle #
TL telomere length
TPOAb thyroid peroxidase antibody
TRI Toxics Release Inventory
TSH thyroid-stimulating hormone
TSLP thymic stromal lymphopoietin
tT total testosterone
tT3 total triiodothyronine
tT4 total thyroxine
TV testicular volume
UCB umbilical cord blood
WAZ weight for age Z-score
WC waist circumference
wk week(s)
WHEALS Wayne County Health, Environment,
Allergy and Asthma Longitudinal
Study
WHO World Health Organization
yr year(s)
8-viii
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APPENDIX 8 REPRODUCTIVE AND
DEVELOPMENTAL EFFECTS
Summary of Causality Determinations for Pb Exposure and
Reproductive and Developmental Effects
This appendix characterizes the scientific evidence that supports causality
determinations for Pb exposure and reproductive and developmental effects. The types
of studies evaluated within this appendix are consistent with the overall scope of the
ISA as detailed in the Process Appendix (see Section 12.4). In assessing the overall
evidence, the strengths and limitations of individual studies were evaluated based on
scientific considerations detailed in Table 12-5 of the Process Appendix
(Section 12.6.1). More details on the causal framework used to reach these conclusions
are included in the Preamble to the ISA (U.S. EPA. 2015). The evidence presented
throughout this chapter supports the following causality conclusions:
Outcome Group
Causality Determination
Pregnancy and Birth Outcomes
Likely to be Causal
Development
Causal
Female Reproductive Function
Likely to be Causal
Male Reproductive Function
Causal
The Executive Summary, Integrated Synthesis, and all other appendices of this Pb
ISA can be found at https://assessments.epa.gov/isa/document/&deid=359536.
8.1 Introduction and Summary of the 2013 Integrated Science
Assessment
This appendix evaluates the epidemiologic and toxicological literature related to the potential
effects of lead (Pb) on reproductive and developmental outcomes, divided into four sections: (1) effects
on pregnancy and birth outcomes; (2) effects on development; (3) effects on female reproductive
function; and (4) effects on male reproductive function. Based on the epidemiologic and toxicological
studies reviewed in the 2013 Pb Integrated Science Assessment (ISA) (U.S. EPA. 2013). the
determination for effects on pregnancy and birth outcomes was based on the mix of inconsistent results of
the epidemiologic and toxicological studies on various birth outcomes, but with some associations
observed in some epidemiologic studies of preterm birth and low birth weight and fetal growth. The
8-1
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determination for developmental effects was informed by evidence from toxicological studies reporting
delayed female sexual maturity and supported by epidemiologic studies of delayed pubertal onset for both
girls and boys. The determination for effects on female reproductive effects was based on epidemiologic
and toxicological studies for reproductive function among females reviewed including endpoints of
hormone levels, fertility, estrous cycle changes, and morphology or histology of female reproductive
organs including the placenta. Of the epidemiologic and toxicological studies reviewed for effects on
female reproductive function, the studies were high-quality and well-designed and examined different
exposure periods in conjunction with a number of outcomes related to female reproductive effects. The
determination for effects on male reproductive function were based on strong toxicological evidence that
showed detrimental effects on semen quality, sperm, and fecundity/fertility, with supporting evidence in
epidemiologic studies of associations between Pb exposure and detrimental effects on sperm. The
summary of the determinations from the 2013 Pb ISA is detailed below.
8.1.1 Effects on Pregnancy and Birth Outcomes
The 2013 Pb ISA (U.S. EPA. 2013) reported the associations between Pb exposure and birth
outcomes (infant mortality and embryogenesis; birth defects; preterm birth; and low birth weight/fetal
growth) were inconsistent overall. There were some associations observed between Pb and low birth
weight when epidemiologic studies used measures of postpartum maternal bone Pb or air exposures. The
associations were less consistent for maternal blood Pb measured during pregnancy or at delivery or
umbilical cord and placenta Pb (maternal blood Pb or umbilical cord and placenta Pb were the biomarkers
most commonly used in studies of low birth weight) but some associations between increased Pb
biomarker levels and decreased low birth weight/fetal growth were observed. Animal studies
investigating the effects of Pb exposure during gestation on litter size, implantation, and birth weight had
varying results between studies. Based on the mix of inconsistent results of studies on various birth
outcomes but some associations observed in select epidemiologic studies of preterm birth and low birth
weight/fetal growth, the evidence in the 2013 Pb ISA was suggestive of a causal relationship between Pb
exposure and birth outcomes.
8.1.2 Effects on Development
The 2013 Pb ISA (U.S. EPA. 2013) reported Pb associated effects on development in
epidemiologic and toxicological studies. Previous toxicological studies indicated that delayed pubertal
onset may be one of the more sensitive developmental effects of Pb exposure with effects observed after
gestational exposures leading to blood Pb levels (BLLs) in the female pup of 1.3-13 (ig/dL (Iavicoli et al..
2006; Iavicoli et al.. 2004). Toxicological studies have reported delayed male sexual maturity as
measured with sex organ weight, seeing significant decrements at BLLs of 20-34 (ig/dL (Ronis et al..
1998c; Sokol et al.. 1985). The 2013 Pb ISA also presented findings from a toxicological study that
8-2
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suggests Pb may act through disruption of insulin-like growth factor 1 (IGF-1) to delay the onset of
puberty, demonstrated by the attenuation of Pb-induced delays in pubertal onset in female rats
supplemented with IGF-1 (Pine et al.. 2006). Thus, data from the toxicological literature and from
epidemiologic studies demonstrated that puberty onset in both males and females is delayed with Pb
exposure. Findings from epidemiologic studies of the effect of Pb on postnatal growth were inconsistent
and findings from the toxicological literature of the effect of Pb exposure were mixed with recent growth
findings showing adult-onset male obesity after gestational and lactational Pb exposure. The 2013 Pb ISA
concluded that, based on the findings of delayed pubertal onset among males and females, there was
sufficient evidence to conclude a causal relationship between Pb exposure and developmental effects.
8.1.3 Effects on Female Reproductive Function
The 2013 Pb ISA (U.S. EPA. 2013) found some evidence of a potential relationship between Pb
exposure and female fertility; however, findings were inconsistent. Epidemiologic studies were largely
cross-sectional and adjustments for important confounding factors were not included in all studies. Some
toxicological studies reported effects on placental pathology and inflammation, decreased ovarian
antioxidant capacity, and altered hormone levels. Overall, the relationship observed with female
reproductive outcomes, such as fertility, placental pathology, and hormone levels in some epidemiologic
and toxicological studies was sufficient for the 2013 Pb ISA to conclude that evidence was suggestive of
a causal relationship between Pb exposure and female reproductive function.
8.1.4 Effects on Male Reproductive Function
The 2013 Pb ISA (U.S. EPA. 2013) reported multiple studies in rodents and non-human primates
that observed Pb-induced sperm DNA damage, reduced sperm quality, reduced sperm production, and
histological and ultrastructural damage to male reproductive organs. Other toxicological studies reported
that Pb exposure was associated with decreases in reproductive organ weights, histological changes in the
testes and germ cell, and subfecundity. The 2013 Pb ISA also presented toxicological evidence suggesting
that Pb may damage sperm cells and sex organ tissue through induction of oxidative stress (Salawu et al..
2009; Shan et al.. 2009; Madhavi et al.. 2007; Rubio et al.. 2006; Wang et al.. 2006). Specifically, one
study reported Pb-induced increases in oxidative stress markers and reductions of levels of antioxidant
enzymes in testicular plasma (Salawu et al.. 2009). In addition, several studies reported attenuation of Pb-
induced reductions in sperm count, motility, and viability when animals were co-administered substances
with known antioxidant properties (Salawu et al.. 2009; Shan et al.. 2009; Madhavi et al.. 2007; Rubio et
al.. 2006; Wang et al.. 2006). Epidemiologic studies were limited due to lack of consideration of potential
confounding factors or the use of men attending a fertility clinic, which could result in a biased sample.
However, a well-conducted epidemiologic study that enrolled men going to a clinic for either infertility
issues or to make a semen donation and controlled for other metals as well as smoking reported a positive
8-3
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association with various detrimental effects in sperm (Telisman et al.. 2007). Studies in the 2013 Pb ISA
that investigated the effects of Pb on hormone levels reported inconsistent results, resulting in uncertainty
as to whether Pb exerts its toxic effects on the reproductive system by affecting the responsiveness of the
hypothalamic-pituitary-gonad axis, by suppressing circulating hormone levels, or by some other pathway.
Based on the consistency and coherence of findings of the detrimental effects of Pb exposure on sperm
and semen in the toxicological literature from animal studies, the support from epidemiologic studies, and
biological plausibility provided by mode of action evidence; however, the evidence in the 2013 Pb ISA
was sufficient to conclude a causal relationship between Pb exposures and male reproductive function.
8.2 Scope
The scope of this section is defined by Population, Exposure, Comparison, Outcome, and Study
Design (PECOS) statements. The PECOS statements define the objectives of the review and establishes
study inclusion criteria thereby facilitating identification of the most relevant literature to inform the Pb
ISA.1 In order to identify the most relevant literature, the body of evidence from the 2013 Pb ISA was
considered in the development of the PECOS statements for this Appendix. Specifically, well-established
areas of research; gaps in the literature; and inherent uncertainties in specific populations, exposure
metrics, comparison groups, and study designs identified in the 2013 Pb ISA inform the scope of this
Appendix. The 2013 Pb ISA used different inclusion criteria than the current ISA, and the studies
referenced therein often do not meet the current PECOS criteria (e.g., due to higher or unreported
biomarker levels). Studies that were included in the 2013 Pb ISA, including many that do not meet the
current PECOS criteria, are discussed in this appendix to establish the state of the evidence prior to this
assessment. With exception of supporting evidence used to demonstrate the biological plausibility of Pb-
associated effects on reproductive and developmental health, studies evaluated and subsequently
discussed within this section were only included if they satisfied all the components of the following
discipline-specific PECOS statement:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect.
Exposure: Exposure to Pb2 as indicated by biological measurements of Pb in the body - with a
'The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
2Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area that was of
particular relevance to the National Ambient Air Quality Standards (NAAQS) review (e.g., longitudinal studies
designed to examine recent versus historical Pb exposure).
8-4
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specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure;3 or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles).
Outcome: Reproductive effects, including but not limited to altered age of puberty onset, reduced
fertility, poor semen quality/motility, and miscarriage. Developmental effects including but
not limited to adverse pregnancy outcomes (e.g., reduced fetal growth, preterm birth, small
for gestational age [SGA], birth defects), as well as postnatal developmental effects.
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (i.e., mouse, rat, Guinea pig,
minipig, rabbit, cat, dog; whole organism) of any lifestage (including preconception, in utero,
lactation, peripubertal, and adult stages).
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a BLL of 30 (ig/dL or below.4'5
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control.
Outcomes: Reproductive and developmental effects.
Study design: Controlled exposure studies of animals in vivo.
8.3 Effects on Pregnancy and Birth Outcomes
The 2013 Pb ISA reported inconsistent findings in the epidemiologic and toxicological literature
for birth outcomes (infant mortality and embryogenesis; birth defects; preterm birth; and low birth
weight/fetal growth). Among the epidemiologic studies, there were inconsistent associations between Pb
3Studies that estimate Pb exposure by measuring Pb concentrations in PMio and PM2.5 ambient air samples are only
considered for inclusion if they also include a relevant biomarker of exposure. Given that size distribution data for
Pb-PM are fairly limited, it is difficult to assess the representativeness of these concentrations to population
exposure [Section 2.5.3 (U.S. EPA. 2013)1.. Moreover, data illustrating the relationships of Pb-PMio and Pb-PNL 5
with BLLs are lacking.
4Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
5This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLLs.
The 95th percentile of the 2011-2016 National Health and Nutrition Examination Survey (NHANES) distribution of
BLL in children (1-5 years; n = 2,321) is 2.66 (ig/dL (CDC, 2019) and the proportion of individuals with BLLs that
exceed this concentration varies depending on factors including (but not limited to) housing age, geographic region,
and a child's age, sex and nutritional status.
8-5
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exposure and preterm birth. A single study of neural tube defects (NTDs) found no associations in the
2013 Pb ISA, but studies within the 2006 Air Quality Criteria Document for Lead (Pb AQCD) (U.S. EPA.
2006) reported associations between Pb exposure and NTDs. There were some associations reported
between Pb and low birth weight when epidemiologic studies used measures of postpartum maternal bone
Pb or air exposures. There were less consistent associations for maternal blood Pb measured during
pregnancy or at delivery or umbilical cord and placenta Pb (maternal blood Pb or umbilical cord and
placenta Pb were the biomarkers most commonly used in studies of low birth weight). The effects of Pb
exposure during gestation in animal toxicological studies included mixed findings, but most studies
reported reductions in birth weight of pups or birth weight of litters when dams were treated with Pb.
The recent epidemiologic and toxicological studies are detailed in the following sections. Effects
on pregnancy and birth outcomes encompass a large range of outcomes. The following sections relating to
pregnancy and outcomes are categorized into seven main sections: (1) maternal health during pregnancy;
(2) prenatal growth; (3) preterm birth; (4) birth defects; (5) spontaneous abortion and pregnancy loss and
fetal and infant mortality; (6) placental function; and (7) other pregnancy and birth outcomes.
8.3.1 Maternal Health During Pregnancy
Maternal health during pregnancy encompasses a wide range of health effects. The details of the
recent epidemiologic and toxicological studies evaluating the association between Pb exposure and
maternal health during pregnancy are provided in Table 8-2 and Table 8-3, respectively.
8.3.1.1 Epidemiologic Studies on Maternal Health During Pregnancy
The main maternal health outcomes evaluated in this section are gestational diabetes mellitus
(GDM) and epigenetic studies. Although there are a limited number of epigenetic studies, these studies
may help to add support for biological plausible pathways for which Pb exposure may affect maternal
health during pregnancy.
8.3.1.1.1 Epidemiologic Studies on Gestational Diabetes Mellitus
There were no studies on GDM in the 2013 Pb ISA. There were several recent epidemiologic
studies that evaluated the association between Pb exposure and GDM and/or impaired glucose tolerance
(IGT) (Tatsuta et al.. 2022a; Zheng et al.. 2021; Zhou et al.. 2021b; Oguri et al.. 2019; Soomro et al..
2019; Wang et al.. 2019; Shapiro et al.. 2015). Generally, across the studies there were null associations
between Pb exposure and GDM and IGT, and/or GDM or IGT. In studies that evaluated Pb in maternal
blood with GDM outcomes, the timing of when Pb was measured differed between trimesters, but the
difference in what trimester Pb was measured did not impact the associations (Oguri et al.. 2019; Soomro
8-6
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et al.. 2019; Wang et al.. 2019; Shapiro et al.. 2015). These studies all reported median BLLs less than
5 (ig/dL (range: 1.7-2.8 (ig/dL) or geometric mean BLLs less than 5 (ig/dL (range: 0.6-1.62 (ig/dL or
6.05-6.13 ng/g). Additionally, while maternal blood was the primary biomarker used to measure Pb
exposure, some studies have used other biomarkers such as maternal serum (Zhou et al.. 2021b) and
maternal erythrocytes (Zheng et al.. 2021); however, the type of biomarker measurement did not
influence the pattern of associations. Only one study reported a decrease of 0.5 (95% confidence interval
[CI]: -1.6, -0.6) mg/dL difference in mid-gestational glucose concentration associated with an
interquartile range (IQR) (17.6 ng/g) change in blood erythrocyte Pb exposure (Zheng et al.. 2021).
Furthermore, multiple studies also considered co-exposure to other metals in addition to Pb, but the
associations remained null (Zheng et al.. 2021; Zhou et al.. 2021b; Oguri et al.. 2019; Wang et al.. 2019).
Overall, the associations between Pb exposure and GDM, IGT, and GDM or IGT were null, and the null
associations persisted across the different trimesters of when Pb levels were measured, the different
biomarkers for Pb exposure, and adjustment for co-exposure to other metals.
8.3.1.1.2 Epidemiologic Studies on Epigenetic Effects During Pregnancy
There were no studies on epigenetic effects during pregnancy evaluated in the 2013 Pb ISA. The
recent epidemiologic studies on epigenetic effects during pregnancy are limited but provide insight on
potential mechanistic pathways in which Pb exposure may impact pregnancy. A single study by Sanders
et al. (2015) assessed the association between maternal Pb levels in blood, patella, and tibial bone and
altered micro RNA (miRNA) expression in the cervix during the second trimester of pregnancy in a
subset of 60 women enrolled in a prospective birth cohort, Programming Research in Obesity, Growth,
Environment and Social Stressors (PROGRESS), in Mexico City. Changes in cervical miRNA expression
are a potential mechanism that could alter gene expression leading to aberrant changes in cervix tissue
function and subsequently impact parturition (Sanders et al.. 2015). Expression of certain miRNAs in the
cervix during pregnancy have been associated with subsequent gestational age (GA) at delivery (Sanders
et al.. 2015). During mid-pregnancy (16-19 weeks gestation), samples from cervical exams were
collected and analyzed for the expression profiles of 800 miRNAs. Overall, there were distinct miRNAs
measured in cervical samples during pregnancy that are associated with the subsequent GA of offspring.
Sanders et al. (2015) also identified differentially expressed miRNAs with respect to preterm compared
term birth in a subset of women. There were two miRNAs expressed in the cervix that were identified in
association with maternal second trimester BLLs, seven miRNAs that were identified in association with
maternal patella bone Pb levels, and six miRNAs that were identified in association with maternal tibia Pb
levels (see Table 8-2). In another epigenetics study in the same PROGRESS cohort, Sanchez-Guerra et al.
(2019) assessed the association of blood Pb exposure during pregnancy with mitochondrial DNA
(mtDNA) content, which is a sensitive marker of mitochondrial function and oxidative stress, in cord
blood. Maternal blood Pb samples were obtained at three time points (second trimester n = 410, third
trimester n = 356, and at delivery n = 354), and cord blood (n = 346) Pb samples were obtained at
delivery. Maternal Pb levels during the second trimester (|3: 0.017 [95% CI: 0.002, 0.031]) were
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associated with higher mtDNA content; however, there were null associations between cord BLLs at
delivery (|3: 0.016 [95% CI: 0.001,0.03]), maternal third trimester blood Pb (|3: 0.015 [95% CI 0.00,
0.03]), and maternal BLLs at delivery (|3: 0.013 [95% CI: -0.001, 0.027]). These epigenetic studies
provide support of potential mechanistic pathways in which Pb exposure is associated with maternal
health during pregnancy.
8.3.1.1.3 Epidemiologic Studies on Other Outcomes Related to Maternal Health During
Pregnancy
There were several other outcomes related to maternal health during pregnancy. More specific
study details, including Pb levels, study population characteristics, potential confounders, and select
results from these studies are highlighted in Table 8-2. In other outcomes related to maternal health
during pregnancy, Pb exposure has been associated with decreased free thyroxine (FT4) during mid-
pregnancy (Kahn et al.. 2014); increased thyroid peroxidase antibodies (TPOAb) during mid-pregnancy
(Kahn et al.. 2014); small increases in umbilical cord blood Pb and elevations in systolic blood pressure
and diastolic blood pressure during labor and delivery (Wells et al.. 2011); changes in Global Severity
Index (GSI), depression and anxiety symptom scores (Li et al.. 2017b); bone mineral density of the
patella (Osorio-Yanez et al.. 2021); increased matrix metalloproteinases (MMP), regulators of uterine
remodeling (Kim et al.. 2022); and increased risk of preeclampsia (Gaiewska et al.. 2021; Wu et al..
2021). However, there were no associations between Pb exposure and reduced Cortisol awakening
response (Braun et al.. 2014); maternal depression (Ishitsuka et al.. 2020); anti-Miillerian hormone
(AMH), a suggested marker of ovarian function and biological marker of female fecundity (Christensen et
al.. 2016); hormone levels in pregnancy (Gustin et al.. 2021); or thyroid function (Corrales Vargas et al..
2022).
8.3.1.2 Toxicological Studies on Maternal Health During Pregnancy
Previous Pb ISAs and AQCDs did not report any toxicological studies that investigated the
effects of Pb on maternal health during pregnancy. Despite this lack of prior studies to compare to, recent
toxicological studies have reported on the effects of Pb on maternal weight gain during pregnancy
(Table 8-3). Maternal weight gain is often used as an indicator of fetal growth and maternal overt toxicity.
Additionally, maternal weight gain shares associations with gestational conditions in humans (Santos et
al.. 2019). Recent studies dosed Sprague-Dawley rats with Pb viagavage for the first 20 days of
pregnancy and reported that the 160 ppm Pb treatment group exhibited reduced weight gain during
pregnancy (maternal BLLs on gestational day [GD] 20 were reported to be 23.9-27.7 (ig/dL) (Saleh et al..
2019; Saleh et al.. 2018). Of note is that both studies by Saleh et al. (Saleh et al.. 2019; Saleh et al.. 2018)
reported reduced brain weight of dams, indicating that overt toxicity may have contributed to the overall
reduction in maternal weight. Further, the reported maternal BLLs were higher than those observed in the
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following studies that investigated the same outcomes. Corv-Slechta et al. (2013) and Schneider et al.
(2016) both dosed C57BL/6 mice with 100 ppm Pb via drinking water starting 2 months prior to mating
and reported no effects on maternal bodyweight gain and observed much lower maternal BLLs with
Corv-Slechta et al. (2013) reporting 12.12 (ig/dL at weaning and Schneider et al. (2016) reporting
12.61 (ig/dL on lactation day 21. Similarly, Wang et al. (2014) reported no effects in Wistar rats dosed
with Pb via drinking water for various durations during pregnancy. In Wang et al. (2014) dosing from
GD 1-10, GD 11-20, and GD 1-20 resulted in maternal BLLs of 26.4, 12.4, and 36.0 (ig/dL, respectively,
at termination of the study on GD 20. Although some of these BLLs overlap with those seen in the studies
by Saleh et al. (2018) and Saleh et al. (2019) wherein suppression of maternal weight gain was observed,
it is possible that the use of different strains or different dosing routes could be attributed to the observed
difference in effect on maternal weight gain. Additionally, the reduction of brain weight observed in the
dams used in the studies by Saleh et al. (2018) and Saleh et al. (2019) suggest that maternal overt toxicity
may be responsible for the observed reduction in maternal weight gain.
8.3.1.3 Integrated Summary of Effects on Maternal Health During Pregnancy
The 2013 Pb ISA did not include epidemiologic and/or toxicological studies that evaluated the
relationship between Pb exposure and maternal health during pregnancy. There were consistent null
associations between Pb exposure and GDM among the recent epidemiologic studies. While the critical
window for GDM is unknown, these studies had different time points during pregnancy in which Pb
exposure was measured and different biomarkers of exposure (blood, serum, and erythrocyte) and the null
associations persisted. A few of the studies were limited by the cross-sectional study design and the small
number of GDM cases. Additionally, a few of the recent epidemiologic studies incorporated mixture
methods to consider Pb exposure in conjunction with co-exposure to other metals to evaluate associations
with GDM, which helps to reduce uncertainties regarding co-pollutant confounding. The limited number
of epigenetic studies provide support of potential mechanistic pathways in which Pb exposure are
associated with selected maternal health during pregnancy. Furthermore, there was a small body of
evidence across various additional pregnancy-related endpoints in the epidemiologic literature; however,
the small number limits the ability to judge coherence and consistency across these studies, although the
positive associations observed demonstrate that Pb exposure could result in physiological responses that
contribute to adverse pregnancy outcomes (e.g., changes in thyroid function, maternal mental health,
changes in blood pressure, preeclampsia). In the recent toxicological literature, there were a limited
number of studies that investigated the relationship between Pb exposure and maternal weight gain during
pregnancy; however, the only studies that observed changes in maternal weight gain also reported signs of
possible overt toxicity (reduced brain weight), indicating that weight gain during pregnancy may not have
been a direct effect of Pb exposure. The majority of recent toxicological studies in rodents reported that
maternal weight gain during pregnancy was unaffected by Pb exposure.
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8.3.2
Prenatal Growth
The recent epidemiologic and toxicological studies that examined the relationship between Pb
exposure and prenatal growth, which includes outcomes such as fetal growth, birth weight, body length at
birth, and GA, are summarized in the text below. Study details of the recent epidemiologic studies are
included in Table 8-4 and the recent toxicological studies are in Table 8-3.
8.3.2.1 Epidemiologic Studies on Prenatal Growth
The epidemiologic studies in the 2013 Pb ISA reported associations between maternal bone Pb
and low birth weight and with studies of Pb air exposures and birth weight. The associations were less
consistent when using maternal blood Pb or umbilical cord and placenta Pb as the exposure measurement,
although some studies did demonstrate associations. The studies of Pb exposure and fetal growth were
limited by their cross-sectional study design, small sample size, high air Pb concentrations (air Pb as high
as 30 (ig/m3), and in some studies, the lack of control of confounders.
A large number of epidemiologic studies have been published since the 2013 Pb ISA on exposure
to Pb and prenatal growth. The studies in this section focus on these prenatal growth outcomes, including
birth weight; low birth weight; body length, crown-to-heel length, head circumference (HC), Ponderal
Index (PI; weight/height3), GA, SGA, and large for gestational age (LGA). Multiple cross-sectional and
cohort studies have been conducted that examined the relationship between Pb exposure and prenatal
growth; however, the findings from the recent epidemiologic studies are inconsistent. There are
differences in study design, timing of the exposure (at different points during pregnancy, at delivery),
differences in biomarkers examined for Pb (maternal blood, maternal serum, cord blood, maternal red
blood cells, placental tissue), and small sample sizes in some studies. The study details, including
information on study population, biomarker of exposure, and outcome, are in Table 8-4.
Several studies used cord blood to assess Pb exposure and reported null associations with birth
weight (Lee et al.. 2021; Govarts et al.. 2020; Tatsuta et al.. 2017; Wang et al.. 2017b; Govarts et al..
2016; Garcia-Esquinas et al.. 2013; Xie et al.. 2013). while a single study reported a reduction in birth
weight (Xu et al.. 2012). Among these studies, there were also inconsistent associations when examining
cord blood Pb exposure and birth weight among infant sex. Tatsuta et al. (2017) evaluated the
associations between cord blood Pb and birth weight between male and female infants, but the
associations remained null. While there were null associations with birth weight, birth length, HC, and PI
when infant sexes were analyzed together, analyses stratifying by infant sex reported associations in male
infants, including increased birth weight (|3: 206.50 [95% CI: 46.15, 366.86]) and decreased HC (|3: -0.65
[95% CI: -1.24, -0.06]) per 1-unit increase logio-Pb cord blood concentration. Among female infants,
there was only a reduction in PI (|3: -0.16 [95% CI: -0.30, -0.02]) per 1-unit increase in the logio-Pb cord
blood concentration (Wang et al.. 2017b).
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In addition to birth weight, there were several other prenatal growth outcomes in these studies
that were evaluated in association with cord blood Pb; however, the associations were inconsistent.
Garcia-Esquinas et al. (2013) also reported null associations between cord blood Pb and birth length, and
1- and 5-minute Apgar scores from the 144 newborns who were a part of cross-sectional biomonitoring
study of the BioMadrid Project. Xu et al. (2012) also reported decreased mean GA of 0.57 weeks (95%
CI: 0.51, 0.63), with increased risk of low birth weight rate (OR: 1.61 [95% CI: 1.37, 1.90]), and
increased risk of intrauterine growth retardation rate (OR: 2.12 [95% CI: 1.68, 2.69]). Xie et al. (2013)
reported a negative association with birth length (|3: -0.84 cm [95% CI: -1.52, -0.16]) per square root 1-
(ig/dL increase in cord blood Pb, but null associations with birth weight (|3: -99.33 g [95% CI: -217.33,
20.67]) and HC (|3: -0.36 [95% CI: -0.81, 0.03]). A single study that was conducted among 1,578
mother-infant pairs in Saudi Arabia reported no associations between cord BLLs and PI below the 10th
percentile (OR: 0.66 [95% CI: 0.42, 1.05]) (Al-Saleh et al.. 2014).
Maternal blood was also used to measure Pb exposure in association with prenatal growth
outcomes in multiple cross-sectional studies, but the associations were inconsistent. Xie et al. (2013)
reported a negative association with birth weight (|3: -148.99g [95% CI: -286.33,-11.66]) per square
root 1 - jag/dL increase in maternal blood Pb measured at delivery, but null associations with birth length
(|3: -0.46 cm [95% CI: -1.25, 0.34]) and HC (|3: -0.37 cm [95% CI: -0.78, 0.19]) among 252 mother-
infant pairs in a rural area located on the south coast of Laizhou Bay, China between 2010 and 2011.
However, Kim et al. (2020) reported negative associations between maternal blood natural log (ln)-Pb,
measured at delivery, and HC (|3: -0.75 cm [95% CI: -1.17, -0.32]) and PI (|3: -0.62 kg/m3 [95% CI:
-1.13, -0.11]), but there were null associations with birth weight (|3: 60 g [95% CI: -15, 135]), BMI (|3:
-0.14 kg/m2 [95% CI: -0.39, 0.11]), and SGA (OR: 0.69 [95% CI: 0.33, 1.46]) among participants of e-
waste Recycling Exposure and Community Health (e-REACH) Study. A study by Xu et al. (2022b)
reported that a one ln-unit increase in maternal BLLs, measured at delivery, was associated with increased
GA (|3: 0.18 weeks [95% CI: 0.05, 0.31]), decreased birth length (|3: -0.39 cm [95% CI: -0.66, -0.22]),
and decreased HC (|3: -0.22 cm [95% CI: -0.39, -0.06]), but a null association with birth weight. There
were also null associations across tertiles of maternal BLLs and low birth weight.
In addition to cord blood and maternal blood, other biomarkers such as maternal serum, cord
blood serum, and placental tissue were used to assess Pb exposure with birth weight among other studies
and reported inconsistent associations (Yang et al.. 2020; Freire et al.. 2019; Mikelson et al.. 2019; Tang
et al.. 2016; Hu et al.. 2015). A study that measured Pb in both maternal serum, measured at delivery, and
cord blood serum reported null associations with birth weight for both biomarkers (Hu et al.. 2015). while
another study reported null associations between cord blood serum and birth weight-for-gestational-age
Z-score, when modeled continuously or categorized by quintiles (Yang et al.. 2020). Although there were
null associations with birth weight and GA, there was a decrease in birth height and a decrease in HC per
ln-Pb increase in umbilical cord serum among 103 mother-newborn pairs from an island in the East China
Sea (Tang et al.. 2016). When placental tissue was the biomarker of exposure for Pb, a single cross-
sectional study reported null associations with birth weight, low birth weight, birth, head, GA, and SGA
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(Frcirc et al.. 2019). but another reported a decrease in birth weight of 58.3 g (95% CI: -97.9, -18.8) per
ln-Pb increase in placental tissue (Mikelson et al.. 2019).
The use of advanced statistical methods to evaluate the impact of co-exposure to other metals, or
mixtures, helps to address uncertainties of co-pollutant confounding. To assess the associations between
metal mixtures (arsenic [As], cadmium [Cd], manganese [Mn], and Pb) in umbilical cord blood and birth
weight, birth length, and HC, 1,088 participants of a birth cohort in Bangladesh were assessed in a cohort
study (Lee et al.. 2021). There were null associations with birth weight (|3: -0.04 g [95% CI: -0.19,
0.11]), birth length (|3: -0.06 cm [95% CI: -0.20, 0.09]), and HC (|3: 0.08 cm [95% CI: -0.06, 0.23]) in
association with an IQR increase in ln-Pb cord blood concentrations, when adjusted for confounders and
other metals. In addition to the multivariable regression analysis, Lee et al. (2021) also used Bayesian
kernel machine regression (BKMR) to estimate the effects of co-exposure to metal mixtures. BKMR is a
method that estimates the multivariable exposure-response function in a flexible and parsimonious way,
conducts variable selection on the (potentially high-dimensional) vector of exposures, and allows for a
grouped variable selection approach that can accommodate highly correlated exposures. In the BKMR
analysis, there was an inverse association between the metal mixture overall and birth length when all
four metal concentrations were >60th percentile and HC when all four metals were >5 5th percentile,
compared to their median values, with stronger associations as the concentrations of the four metals
increased. However, when estimating the difference in birth size with an IQR increase in each individual
metal when the other metals were fixed at their 25th, 50th, or 75th percentiles, the associations with Pb
were null.
Overall, in the multiple longitudinal birth cohort studies, there were inconsistent findings between
various Pb exposure biomarkers and prenatal growth outcomes. The multiple longitudinal birth cohort
studies have reported inconsistent associations. These studies collected maternal samples during different
time periods during pregnancy and utilized different biomarkers to measure Pb exposure to evaluate
associations with a variety of prenatal growth outcomes. In the longitudinal studies that measured Pb
exposure from maternal blood, there were inconsistent patterns of association with prenatal growth
outcomes, regardless of the trimester Pb exposure was measured or prenatal growth outcome (see
Table 8-4). Several studies reported null associations with birth weight (Shih et al.. 2021; Woods et al..
2017; Taylor et al.. 2016; Bloom et al.. 2015; Garcia-Esquinas et al.. 2014; Rabito et al.. 2014) and birth
weight Z-score (BWZ) (Daniali et al.. 2023). while others reported reductions in birth weight (Goto et al..
2021; Hu et al.. 2021; Rodosthenous et al.. 2017; Taylor et al.. 2015). Of note, Rodosthenous et al. (2017)
measured Pb levels in maternal blood during the second trimester among 944 mother-infant pairs in the
PROGRESS cohort in association with birth weight using both linear and quantile regression. While the
linear regression reported a null association with birth weight-for-gestational-age Z-score (|3: -0.06 [95%
CI: -0.13, 0.003]) per log2-Pb blood level increase, the quantile regression analysis revealed larger
magnitudes of the association maternal blood Pb and birth weight-for-gestational-age Z-score. The
magnitude of the association was largest in the lowest (<30th) Z-score percentiles (difference in Z-score
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ranged from -0.13 to -0.08). The use of quantile regression provides insights to potential sensitivity to Pb
exposure for smaller infants, an association that was not detected by linear regression.
While some studies reported null associations with birth length (Daniali et al.. 2023; Shih et al..
2021; Bloom et al.. 2015). a single study reported a 0.20 cm decrease (95% CI: -0.30, -0.10) in birth
length per 1 (ig/dL increase in maternal BLL (collected during the second or third trimester) among
participants of the Japan Environment and Children's Study (JECS) (Goto et al.. 2021). A single study
reported a reduction in HC of 0.03 cm (95% CI -0.06, -0.00) per 1 (ig/dL increase of first trimester
maternal blood Pb (Taylor et al.. 2016). but other studies reported null associations with HC and first
trimester maternal blood Pb (Daniali et al.. 2023; Taylor et al.. 2015). pre-pregnancy maternal and
parental blood (Bloom et al.. 2015). or second or third trimester maternal blood Pb (Shih et al.. 2021).
While there was reported decreased GA (|3: -1.9 days [95% CI: -3.1, -0.5]) per IQR increase in second
trimester maternal ln-Pb blood level among those in Puerto Rico Testsite for Exploring Contamination
Threats (PROTECT) cohort (Ashrap et al.. 2020). there were null associations with gestational and pre-
pregnancy maternal and parental blood (Bloom et al.. 2015) and maternal BLLs during the second or third
trimester among the JECS (Goto et al.. 2021); however, Goto et al. (2021) did report an increased risk of
SGA (OR: 1.34 [95% CI: 1.16, 1.55]) and increased risk of low birth weight (OR: 1.34 [95% CI: 1.16,
1.55]) per 1 (ig/dL increase in maternal BLL, but other studies did not report increased risk of SGA
(Thomas et al.. 2015) and maternal blood (collected during the first and third trimesters of pregnancy) or
second trimester maternal blood (Ashrap et al.. 2020). There were consistent null associations with PI and
maternal blood (Shih et al.. 2021; Bloom et al.. 2015) and crown-to-heel length and first trimester
maternal blood Pb (Taylor et al.. 2016; Taylor et al.. 2015).
In addition, some of the longitudinal studies considered different effect modifiers when assessing
the associations between maternal BLLs and prenatal growth outcomes. Among participants in the
Canadian Maternal-Infant Research on Environmental Chemicals (MIREC) study, there were null
associations between maternal blood (collected during the first and third trimesters of pregnancy) and
SGA across tertiles of maternal blood Pb (Thomas et al.. 2015). In addition, an exploratory analysis was
conducted to examine the potential effect modification of single nucleotide polymorphisms (SNP) in
GSTP1 and GSTOl genes on the relationship of maternal blood Pb and SGA. There was a marginal
interaction between maternal Pb exposure and the GSTP1 Al 14V SNP (p = 0.06), but there was no
indication of effect modification by other GSTP1 and GSTO1 SNPs on the associations between maternal
blood Pb and SGA. In another study in the PROTECT cohort, the modifying effect of psychosocial stress
on the association between maternal blood Pb exposure and GA, BWZ, SGA, and LGA were examined in
a subset of 682 pregnant women (Ashrap et al.. 2021). Maternal blood samples were collected at
18 ± 2 weeks gestation and 26 ± 2 weeks gestation. Among mothers who reported "good" psychosocial
status, there was decreased gestation age (|3: -1.9 days [95% CI: -3.2, -0.6]); however, there were null
associations with BWZ (|3: 0.1 [95% CI: 0.0, 0.2]), SGA (OR: 0.86 [95% CI: 0.65, 1.14]), and LGA (OR:
0.89 [95% CI: 0.64, 1.23]). The associations for mothers who reported "poor' psychosocial status were
null across the birth outcomes.
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In addition to the associations between prenatal growth outcomes and Pb levels, there were sex-
stratified differences. In a study by Garcia-Esquinas et al. (2014). 97 mother-father-infants in the
BioMadrid Study were used to evaluate associations between prenatal Pb exposure and fetal development
from three biomarkers (maternal and paternal blood Pb at 34 weeks gestation and cord blood at delivery)
with different growth metrics at birth. While there were no associations between log-Pb blood levels
(maternal, paternal, or cord) and gestation age, birth weight, birth length, abdominal diameter, or cephalic
diameter (CD), associations were observed when analyses were stratified by infant sex. Among female
infants, there was decreased birth length of 1.06 cm (95% CI: -2.03, -0.08) and CD of-0.55 cm (95%
CI: -1.03, -0.07) per two-fold increase in paternal BLLs (|ig/L). but there were no associations among
male infants. In the study by Shih et al. (2021). there were null associations between maternal blood log2-
Pb concentrations (collected between 6 and 32 weeks of gestation) and prenatal growth outcomes (GA,
birth weight, birth length, HC, and PI). However, when stratified by infant sex, there were reductions in
GA (|3: -0.98 weeks [95% CI: -1.67, -0.30]), birth weight (|3: -381 g [95% CI: -583, -178]), birth length
(|3: -1.44 cm [95% CI: -2.45, -0.42]), and HC (|3: -1.10 cm [95% CI: -1.70, -0.50]), but had a null
association with PI (|3: -1.07 kg/m3 [95% CI: -1.56, 0.39]) among female infants, while the associations
for these same outcomes were null among male infants.
There were also a limited number of studies that considered co-exposure to other pollutants. From
the MIREC study, 1,857 mother-infant pairs were analyzed to examine the relationship between prenatal
exposure to a mixture of endocrine-disrupting chemicals, including Pb, and birth weight using BKMR
(Hu et al.. 2021). Maternal blood was collected during the first trimester of pregnancy. In the adjusted
model for log2-Pb, every two-fold increase in Pb concentration was associated with a mean birth weight
reduction of 82.22 g (95%: -145.46, -18.97), and when adjusted for other metals, the reduction in mean
birth weight was 75.89 g (95% CI: -141.24, -10.54). In the mixtures analysis, Pb was the main
contributor to the adverse effect on birth weight in the metal mixture consisting of As, Cd, mercury (Hg),
Mn, and Pb. An increase in the log2-Pb concentration from the 25th to the 75th percentile was associated
with a posterior mean of -47g, meaning that there was a reduction in mean birth weight of 47 g, while
holding the other components in the metal mixture constant at their median values.
In addition to maternal blood Pb, other biomarkers such as maternal erythrocytes, maternal
serum, and teeth were used to assess Pb exposure with prenatal birth outcomes, including birth weight,
birth length, or HC. Maternal erythrocytes from blood samples were collected during the third trimester
(mean: gestational week 29) from 584 mothers in the Nutritional impact on Immunological maturation
during Childhood in relation to the Environment (NICE) study in Northern Sweden (Gustin et al.. 2020).
Maternal erythrocytes reflect exposure over the past 1-3 months. A doubling of maternal erythrocyte Pb
concentration was not associated with birth weight (|3: -13 g [95% CI: -66, 41]), birth length (|3:
-0.080 cm [95% CI; -0.31, 0.15]), or HC (|3: 0.059 cm [95% CI: -0.22, 0.34] for maternal erythrocyte Pb
concentration less than the median of 14 |ig/kg and |3: -0.24 cm [95% CI: -0.53, 0.056] for maternal
erythrocyte Pb concentration greater than median of 14 |ig/kg). There was no interaction by infant sex.
When mutually adjusted for other maternal metal exposure to Cd and Hg, the null associations persisted.
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In a subset of the Project Viva prospective pre-birth cohort, individual and joint effects of metal mixture
components on birth weight, length, HC, and GA were estimated in association with maternal erythrocyte
Pb concentrations collected during early pregnancy (11.3 ± 2.8 weeks of gestation) from 1,423 mother-
infant pairs (Rahman et al.. 2021). In single metal model, an IQR increase in maternal erythrocyte Pb
concentration was associated with a 33.9 g (95% CI: -65.3, -2.5) decrease in birth weight, but there were
no associations with birth length (|3: -0.10 cm [95% CI: -0.29, 0.09]), HC (|3: -0.07 cm [95% CI: -0.17,
0.04]), or GA (|3: 0.03 weeks [95% CI: -0.10, 0.16]). When stratified by infant sex, the associations were
null for both male and female infants and birth weight, birth length, HC, and GA. Additionally, there was
consistent pattern of association of decreased birth weight, birth length, and HC overall and in the infant
sex-stratified analyses (see Table 8-4).
A total of 3,125 mother-infant pairs were recruited from the China-Anhui Birth Cohort Study
(C-ABCS) to investigate the associations between maternal serum Pb levels the first trimester (median of
11 weeks gestation) and in the second trimester (median of 16 weeks gestation) with growth metrics
(Wang et al.. 2017a). Overall maternal serum Pb during pregnancy had a negative association with birth
weight (|3: -2.74 g [95% CI: -5.17, -0.31]), but null associations with birth length, HC, and chest
circumference. When stratified by trimester, the negative association with birth weight persisted, with a
reduction of 4.40 g (95% CI: -8.22, -0.58) for first trimester maternal serum Pb and a 1.64 g (95% CI:
-4.80, -0.58) reduction for second trimester maternal serum Pb. There were no associations by trimester
maternal serum Pb for birth length, HC, or chest circumference. In addition, there was increased risk of
SGA of 1.45 (95% CI: 1.04, 2.02) for subjects with medium-Pb maternal serum (1.18-1.70 (ig/dL) and
increased risk of SGA of 1.69 (95% CI: 1.22, 2.34) in subjects with high-Pb maternal serum
(>1.71 (ig/dL), compared to low-Pb maternal serum (<1.18 (.ig/dL). When stratified by infant sex, there
was increased risk of SGA among female infants (OR: 1.51 [95% CI: 0.99, 2.31] for medium-Pb maternal
serum and OR: 1.68 [95% CI: 1.12, 2.54] for high-Pb maternal serum), but among male infants, the
associations were null. There was an increased risk of SGA with high first trimester maternal serum Pb
(OR: 2.13 [95% CI: 1.24, 3.38]), but there were null associations among second trimester maternal serum
Pb.
In a small cohort study, second and third trimester Pb levels were estimated from baby teeth from
145 participants in the Wayne County Health, Environment, Allergy and Asthma Longitudinal Study
(WHEALS) (Cassidv-Bushrow et al.. 2019). There were no associations between tooth Pb in the second
or third trimester and BWZ (|3: -0.15 [95% CI: -0.35, 0.05] for second trimester and |3: -0.06 [95% CI:
-0.24, 0.12] for third trimester) or GA at delivery (|3: 0.08 [95% CI: -0.19, 0.35] for second trimester and
|3: 0.14 [95% CI: -0.11, 0.39] for third trimester) in the fully adjusted models. There was no indication
that there was a time effect (difference between the effect estimates in the second and third trimesters) for
birth weight for Z-score (|3: -0.31 [95% CI: -0.90, 0.28]) or GA at delivery (|3: -0.22 [95% CI: -1.08,
0.64]). Additionally, when stratified by child's sex, there were no associations between tooth Pb in the
second or third trimester and BWZ or GA at delivery.
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In a study by Bui et al. (2022). effects of short-term maternal exposure to airborne Pb during
pregnancy on birth weight, low birth weight, and SGA was estimated using a quasi-experimental variation
in airborne Pb exposure based on the National Association for Stock Car Auto Racing (NASCAR)'s
deleading of racing fuel in a difference-in-difference model in the Charlotte-Concord-Gastonia
Metropolitan Statistical Area in North Carolina. After deleading of racing fuel, there was an average
increase in birth weight of 102.50 g (95% CI: 45.73, 159.2), decreased probability of low birth weight of
0.0445 (95% CI: -0.0697, -0.0194), and reduction in the probability of SGA of 0.0396 (95% CI:
-0.0638, -0.0155) among children born to mothers residing less than 4000 meters of the Charlotte Motor
Speedway, compared with those residing greater than 10,000 meters. The difference-in-difference
methodology allows for the control of time-varying confounders, removing biases from comparisons over
time in the treatment group that could be the result of trends due to other causes of the outcome.
8.3.2.2 Toxicological Studies on Prenatal Growth
The 2013 Pb ISA discussed a few studies that reported reduced birth weight of offspring from Pb-
treated dams (Masso-Gonzalez and Antonio-Garcia. 2009; Wang et al.. 2009; Teiion et al.. 2006). Recent
toxicological studies consistently report no effects of Pb on birth weight (Table 8-3). Most studies began
exposure of the dam prior to conception of the offspring (Zhao et al.. 2021; Tartaglionc et al.. 2020; Rao
Barkur and Bairv. 2016; Schneider et al.. 2016; Barkur and Bairv. 2015; Weston et al.. 2014; Corv-
Slechta et al.. 2013) and a few studies began exposure of the dam at the time of conception (GD 0) (Rao
Barkur and Bairv. 2016; Barkur and Bairv. 2015; Barkur et al.. 2011). Of note is that Teiion et al. (2006).
a study discussed in the 2013 Pb ISA, elaborated that the observed reduction in litter weights born to Pb-
treated dams was largely driven by the reduced size of female pups, whereas males were unaffected. In
agreement, some recent studies that reported no effect of Pb on birth weight assessed weight in male pups
only (Barkur and Bairv. 2015; Barkur et al.. 2011). However, all other recent studies included females in
birth weight analyses and reported no effects of Pb on birth weight of exposed offspring.
8.3.2.3 Integrated Summary of Effects on Prenatal Growth
The epidemiologic studies in the 2013 Pb ISA reported associations between maternal bone Pb
and low birth weight and with studies of Pb air exposures and birth weight. The associations were less
consistent when using maternal blood Pb or umbilical cord and placenta Pb as the exposure measurement
although some studies did demonstrate associations. The studies of Pb exposure and fetal growth were
limited by cross-sectional study design, small sample size, high Pb concentrations (air Pb as high as
30 (ig/m3), and in some studies, the lack of control of confounders. A recent quasi-experimental study of
maternal exposure to airborne Pb during pregnancy found an increase in birth weight, decreased
probability of low birth weight, and reduction in the probability of SGA after the deleading of racing fuel.
However, overall, the recent epidemiologic studies reported inconsistent associations between Pb
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exposure and prenatal growth outcomes, while the toxicological studies consistently reported no effects of
Pb on offspring birth weight. The inconsistent findings from the recent epidemiologic studies may be due
to differences in study design, timing of when the exposure was measured (e.g., during pregnancy, at
delivery), biomarkers examined for Pb (e.g., maternal blood, cord blood, maternal red blood cells,
maternal serum, placental tissue), difference in growth metrics assessed (e.g., birth weight, birth length,
GA), and small sample sizes in some studies. While there were inconsistencies in the findings among the
epidemiologic studies, the recent epidemiologic studies were able to address a few of the uncertainties in
the 2013 Pb ISA. Many of the recent studies were conducted in well-designed longitudinal birth cohorts,
considered the differences in effects by infant sex, and controlled for wide range of confounders,
including GA (when not an outcome of interest), and maternal health factors (e.g., smoking, parity, BMI).
Additionally, some epidemiologic studies controlled for other metal exposure, and other studies evaluated
the associations with joint effects or as a mixture. A few toxicological studies were reviewed in the 2013
Pb ISA, all of which reported reductions in birth weight of offspring born from Pb-exposed dams.
However, recent toxicological studies do not support previous studies and consistently report no effects of
Pb on offspring birth weight.
8.3.3 Preterm Birth
The recent epidemiologic and toxicological studies that examined the relationship between Pb
exposure and preterm birth are summarized in the text below. Study details of the recent epidemiologic
studies are included in Table 8-5 and the recent toxicological studies are in Table 8-3.
8.3.3.1 Epidemiologic Studies on Preterm Birth
The epidemiologic studies reviewed in the 2013 Pb ISA reported overall inconsistent findings
regarding a relationship between indicators of Pb exposure and preterm birth. However, there were a few
well-conducted epidemiologic studies that reported associations between maternal blood Pb and preterm
birth (Vigeh et al.. 2011; Jelliffe-Pawlowski et al.. 2006).Among the epidemiologic studies , there were
no apparent patterns within the type of exposure measurement or Pb level. Many of these studies are
limited by the small number of preterm births and their cross-sectional design (i.e., studies of umbilical
cord blood may not adequately characterize BLLs earlier in pregnancy). Among the longitudinal cohort
studies, the results were mixed, with some studies reporting associations between maternal blood Pb
during pregnancy and preterm birth. Most studies controlled for potentially important confounders, such
as maternal age and smoking.
In the recent epidemiologic studies examining the risk of preterm birth and Pb exposure, the
findings were generally consistent (Table 8-5). Most notably is a quasi-experimental study employing
difference-in-difference methodology. In a study by Bui et al. (2022). the effects of short-term maternal
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exposure to airborne Pb during pregnancy on preterm birth was estimated using a quasi-experimental
variation in airborne Pb exposure based on NASCAR" s deleading of racing fuel in a difference-in-
difference model in the CCG MSA in North Carolina. There was decreased probability of preterm birth of
0.295 (95% CI: -0.0572, -0.000185) among children born to mothers residing less than 4000 meters of
the CMS, compared to those residing greater than 10,000 meters after deleading of racing fuel. The
difference-in-difference methodology allows for the control of time-varying confounders, removing
biases from comparisons over time in the treatment group that could be the result of trends due to other
causes of the outcome.
In a cross-sectional study of 696 mother-infant pairs in the Study on the Environment and
Reproductive Health (EMASAR) cohort in Argentina, the relationship between maternal Pb levels, which
were collected 36 ± 12 hours postpartum, and preterm birth was examined (Xu et al.. 2022b). Among
tertiles of maternal Pb levels, there were null associations with preterm birth (OR: 1.24 [95% CI: 0.35,
4.4] in tertile 2 and OR: 1.26 [95% CI: 0.32, 5.00] interfile 3). In another study, cord blood samples were
obtained from 432 infants born in an area with e-recycling (Guiyu) and 99 from an area without e-
recycling (Xiamen) in China, but there was no increased risk of preterm birth (OR: 1.09 [95% CI: 0.93,
1.28]) (Xu et al.. 2012). Additionally, another study used placental tissue Pb levels in association with
preterm birth among 327 mother-infant pairs who were part of the Instituto de Nanociencia y Materiales
de Aragon (INMA) Project in Spain and found no association with risk of preterm birth (OR: 0.40 [95%
CI: 0.04, 4.70]) (Freire et al.. 2019).
In a case-control study, maternal serum Pb, collected during the first or second trimester, was not
associated with risk of spontaneous preterm birth (OR: 1.46 [95% CI: 0.97, 2.18]) among 147 cases and
381 controls (Yu et al.. 2019). When stratified by the trimester of collection of maternal serum Pb, there
was null association for spontaneous preterm birth (OR: 1.63 [95% CI: 0.91, 2.91]) with first trimester
maternal serum Pb only or second trimester maternal serum Pb only (OR: 1.27 [95% CI: 0.71, 2.28]). In a
nested case-control study, the association between exposure to 41 metals/metalloids, including Pb, during
early pregnancy measured in maternal serum and risk of spontaneous preterm birth was investigated (Xu
et al.. 2022a). There were 74 cases of spontaneous preterm birth and 74 controls. In the highest quartile of
maternal serum Pb levels, there was an increased risk of spontaneous preterm birth of 4.09 (95% CI: 1.31,
12.77) and there was evidence of potential exposure-response across the quartiles (p for trend: 0.017).
Tsuii et al. (2018) used data on 14,847 pregnant women who were participants of the JECS to
assess the association between second and third trimester maternal blood (collected at gestational weeks
14-39) and early preterm (22 to <34 weeks) and late preterm (34 to <37 weeks). Among the quartiles of
Pb exposure, there was no increased risk in early preterm birth or late preterm birth. There was also no
evidence of a linear exposure-response trend among the Pb exposure quartiles in either the early preterm
births (p for trend: 0.134) or late preterm (p for trend: 0.920). In another cohort study using data from the
JECS, per each 0.1 (ig/dL increase in maternal BLL, there was no increased risk of preterm delivery (OR:
0.90 [95% CI: 0.70, 1.16]) (Goto et al.. 2021).
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In a small cohort (n = 98) from the Conditions Affecting Neurocognitive Development and
Learning in Early Childhood (CANDLE) study in Shelby County, TN, Pb was measured cord blood and
from maternal blood collected during the second and third trimester, at delivery (Rabito et al.. 2014).
Each 0.1-unit increase in maternal blood Pb in the second trimester (OR: 1.66 [95%CI: 1.23, 2.23]) and
third trimester (OR: 1.24 [95% CI: 1.01, 1.52]) was positively associated with preterm birth, but there was
no increased risk of early-term birth (>37 to <39 weeks) associated with maternal blood Pb in the second
trimester (OR: 0.87 (95% CI: 0.63, 1.20]) and third trimester (OR: 0.88 [95% CI: 0.69, 1.13]).
In the Avon Longitudinal Study of Parents and Children, maternal blood samples were collected
as early as possible in pregnancy, with a median GA of 11 weeks at the time of sampling (range 1-
42 weeks, IQR 9-13 weeks) (Taylor et al.. 2015). There was increased risk of preterm delivery (OR: 2.00
[95% CI: 1.35,3.00]) for maternal BLLs >5 (ig/dL. Li et al. (2017a) investigated the associations between
maternal serum Pb levels and risk of preterm birth in a population-based birth cohort (n = 3,125), part of
the China-Anhui Birth Cohort. Maternal serum Pb levels were categorized into tertiles: low-Pb
(<1.18 (ig/dL), medium-Pb (1.18-1.70 (.ig/dL). and high-Pb (>1.71 (.ig/dL). There was an increased risk of
preterm birth in the medium-Pb tertile (OR: 2.33 [95% CI: 1.49, 3.65]) and high-Pb tertile (OR: 3.09
[95% CI: 2.01,4.76]).
In a study by Ashrap et al. (2020). individual and mixture effects of metals and metalloids on
preterm birth among 731 pregnant women in the PROTECT cohort were examined. Maternal blood was
collected at 16-20 and 24-28 weeks gestation. There was an increased risk of preterm birth (OR: 1.63
[95% CI: 1.17, 2.28]) and spontaneous preterm birth (OR: 1.53 [95% CI: 1.00, 2.35]) per IQR increase in
maternal blood Pb in the individual pollutant model. The mixture pollutant models and elastic net
regularization identified Pb and zinc as the most important predictors of preterm birth, while BKMR
method identified Pb, zinc, and Mn as most predictive of preterm birth. In another study in the PROTECT
cohort, the modifying effect of psychosocial stress on the association between Pb and overall preterm
birth (<37 completed weeks of gestation) and spontaneous preterm birth (<37 completed weeks of
gestation defined as presentation of premature rupture of the membranes, spontaneous preterm labor, or
both) (Ashrap et al.. 2021). There was an increased risk of overall preterm birth among mothers who
reported "good" psychosocial status (OR: 1.72 [95% CI: 1.14, 2.58]), but null association among mothers
who reported "poof' psychosocial status (OR: 1.43 [95% CI: 0.69, 2.97]). There were null associations
among mothers who reported "good" psychosocial status and "poor" psychosocial status and spontaneous
preterm birth (OR: 1.56 [95% CI: 0.93, 2.6] and OR: 1.22 [95% CI: 0.42,3.56], respectively).
8.3.3.2 Toxicological Studies on Preterm Birth
Both the 2013 Pb ISA and the 2006 Pb AQCD did not describe any studies that reported on the
effects of Pb on preterm birth in animals. Only one recent study was found that reports on gestation
duration (Betharia and Maher. 2012). Betharia and Maher (2012) reported no effect of Pb on gestation
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term when Sprague-Dawley rats were dosed from GD 0 to postnatal day (PND) 20. BLLs were measured
in offspring and reported to be 9.03 (ig/dL on PND 2, 0.976 (ig/dL on PND 25, 0.0318 (ig/dL on PND 60.
8.3.3.3 Integrated Summary of Effects on Preterm Birth
In summary, there were inconsistencies in the recent epidemiologic studies examining the
relationship between Pb exposure and risk of preterm birth, similar to the 2013 Pb ISA. There was no
apparent pattern associated with any biomarker of Pb exposure. Several of the recent epidemiologic
studies were conducted in well-designed, longitudinal birth cohorts, and controlled for wide range of
confounders, including GA, other metals, and maternal health factors (e.g., smoking, parity, BMI).
Overall, among the epidemiologic studies, there was a pattern of elevated risk of preterm birth observed
across several studies from multiple geographic locations, including a quasi-experimental study. Among
these studies, there were still some uncertainties in the timing of the exposure (e.g., during pregnancy, at
delivery), and biomarkers examined for Pb (e.g., maternal blood, cord blood, maternal red blood cells,
maternal serum, placental tissue). There were no toxicological studies that investigated the effects of Pb
on preterm birth in the 2013 Pb ISA and the 2006 Pb AQCD, and recent toxicological data are sparse with
only a single PECOS-relevant study available, which reported no effects of Pb on gestation duration,
making it difficult for toxicological data to support epidemiologic evidence.
8.3.4 Birth Defects
The recent epidemiologic and toxicological studies that examined the relationship between Pb
exposure and birth defects are summarized in the text below. Study details of the recent epidemiologic
studies are included in Table 8-6 and the recent toxicological studies are in Table 8-3.
8.3.4.1 Epidemiologic Studies on Birth Defects
In the 2013 Pb ISA, there were only a few studies available for review evaluating associations
between Pb exposure and birth defects, specifically NTDs. These studies did not report associations
between Pb exposure and NTDs. These studies were limited by the timing of Pb measurements, whether
taken at delivery or postnatally, and the lack of potential confounders.
A few recent epidemiologic studies examined the relationship between Pb levels and birth
defects. Several studies evaluated the association between NTDs in different biomarkers (placental tissue,
umbilical cord tissue, and maternal serum) (Liu et al.. 2021; Tian et al.. 2021; Jin et al.. 2013). Recent
studies that measured Pb exposure from placental tissue or umbilical cord tissue reported no increased
risk for NTDs overall or by subtype (Liu et al.. 2021; Jin et al.. 2013) (see Table 8-6 for details).
However, in a case-control study which evaluated the single and joint effects of 10 metals measured in
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maternal serum during pregnancy, there was increased risk for NTDs (Tian et al.. 2021). In the single
pollutant model, there was increased risk of NTD of 2.05 (95% CI: 1.05, 4.02) in the second tertile and
3.51 (95% CI: 1.76, 6.98) in the third tertile, relative to the lowest tertile of maternal serum Pb levels,
indicating an exposure-response relationship (p for trend: <0.001). There was also increased risk by NTD
subtype. There was increased risk of spina bifida of 2.16 (95% CI: 1.00, 4.88) in the second tertile and
5.16 (95% CI: 2.24, 11.87) in the third tertile, relative to the lowest tertile of maternal serum Pb levels,
with an indication of an exposure-response relationship (p for trend: 0.022). For anencephaly, there was
increased risk of 2.97 (95% CI: 1.09, 8.12) in the second tertile and 5.54 (95% CI: 1.89, 16.19) in the
third tertile, relative to the lowest tertile of maternal serum Pb levels, with an indication of an exposure-
response (p for trend: 0.002). Among female infants, there was increased risk of NTD of 6.45 (95% CI:
2.20, 18.95) in the highest tertile, relative to the lowest tertile of maternal serum Pb levels, with exposure-
response relationship (p for trend: 0.001). Among male infants, there was increased risk of NTD of 2.16
(95% CI: 1.03, 4.59), and an indication of an exposure-response relationship (p for trend: 0.048).
Pi et al. (2018) investigated the associations between placental Pb concentrations and the risk of
orofacial cleft (OFC) defects among 103 cases and 206 controls in northern China. With increasing
tertiles of placenta Pb concentrations (p for trend <0.001), there was increased odds of orofacial defects of
3.88 (95% CI: 1.78, 8.42) for those in the second (57.5-96.8 ng/g dry weight) tertile of placenta Pb
exposure and 5.17 (95% CI: 2.37, 11.29) for those in the highest (>96.8 ng/g dry weight) tertile of
placenta Pb exposure, compared to the lowest (<57.5 ng/g dry weight). When restricting to those with
higher than the median placenta Pb concentration (>77.2 ng/g), there was increased risk of 3.08 (95% CI:
1.74, 5.47) of OFC defects among 71 cases and 84 controls. However, in a nested case-control study
among a subset of participants in the JECS, Takeuchi et al. (2022) did not find increased risk of cleft lip
and palate (n = 192 cases and n = 1,920 matched controls) and second trimester maternal blood Pb
concentrations (OR: 1.10 [95% CI: 0.55, 2.21]), which controlled for co-exposure to three other metals
(Hg, Cd, and Mn) in the multivariate model.
In another study of the JECS, maternal serum Pb samples were collected during mid- and late
gestation and were evaluated in association with congenital abdominal malformations (Mivashita et al..
2021). There were 139 cases and 89,134 controls. There were null associations across the quartiles of
maternal serum Pb concentrations and any abdominal malformations, with no exposure-response
relationship across quartiles (p for trend: 0.233). The null associations persisted for the subtypes of
congenital abdominal malformations, but there was an inverse exposure-response relationship observed
across the quartiles of maternal serum Pb and omphalocele (p for trend: 0.033).
A single study explored the associations between umbilical cord serum Pb levels and congenital
heart disease (CHD) birth defects among 97 case and 201 controls (Liu et al.. 2018). In the highest
umbilical serum Pb group (>8.26 ng/mL), the odds of CHD were 1.67 (95% CI: 0.88, 3.17) compared to
the those in the lowest umbilical serum Pb group (<6.69 ng/mL). The odds by CHD subtypes were near
null, (CIs include 1) (see Table 8-6).
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8.3.4.2
Toxicological Studies on Birth Defects
The 2013 Pb ISA did not report any toxicological studies that investigated the effects of Pb on
birth defects. The 2006 Pb AQCD described studies that reported Pb-induced birth defects; however,
these findings were confounded by maternal toxicity (Dev et al.. 2001; Ronis et al.. 1996; Flora and
Tandon. 1987). Two recent studies published since the 2013 Pb ISA have investigated Pb-induced birth
defects in offspring in rodents (Table 8-3). Both studies dosed Wistar rats with 0.2% Pb in the drinking
water for varying duration, including dosing starting 30 days prior to gestation and ending the day prior to
mating, dosing from GD 0 to PND 21, and dosing from GD 0 to 21 (Rao Barkur and Bairv. 2016; Barkur
and Bairv. 2015). Offspring BLLs measured on PND 22 varied between 3.02-3.03 (ig/dL for animals
from dams dosed prior to gestation, 5.30-5.51 (ig/dL for animals from dams dosed during gestation, and
31.6-32.0 (ig/dL for animals from dams dosed from the beginning of gestation to lactation day 21. No
maternal toxicity was apparent in any of the recent studies, suggesting that the contrast found between the
lack of malformations observed in these recent publications and the reported malformations described in
the 2006 Pb AQCD may be attributed to a lack of maternal toxicity due to the use of lower doses in more
recent studies.
8.3.4.3 Integrated Summary of Effects on Birth Defects
The studies reviewed in the 2013 Pb ISA did not report associations between Pb exposure and
NTDs. Among the recent epidemiologic studies, there were inconsistent associations with Pb exposure
and NTDs, congenital heart defects, and OFC defects overall. While the associations were generally null
for NTDs and CHDs, there was a pattern of positive associations with OFC defects. The inconsistencies
in these findings are limited by the different birth defects of interest, the small sample sizes given the rare
outcome, timing of Pb exposure (different measurements to estimate exposure during pregnancy),
differences in the biomarker tested, and the confounders considered in the analyses. The recent
epidemiologic studies controlled for a wide range of potential confounders; however, which was a
limitation from the 2013 Pb ISA. Further, some studies considered co-exposure to other metals and
differences by infant sex. Some previous toxicological studies reported that Pb exposure resulted in birth
defects in offspring, but it was noted that these studies often used doses so high that maternal toxicity
occurred as well. Recent toxicological studies report no effects of Pb on birth defects in offspring and also
do not report that maternal toxicity occurred, further supporting that maternal toxicity may have been
involved with the birth defects observed in previous studies.
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8.3.5 Spontaneous Abortion and Pregnancy Loss and Fetal and Infant
Mortality
The 2013 Pb ISA concluded that the toxicological and epidemiologic data provided inconsistent
findings for the role of Pb in spontaneous abortions, while there were no available epidemiologic or
toxicological studies on the relationship between Pb levels and infant mortality. The recent epidemiologic
and toxicological studies examining the relationship between Pb exposure and spontaneous abortion,
pregnancy loss, and fetal and infant mortality are summarized in the text below. Study details of the
recent epidemiologic studies are included in Table 8-7 and the recent toxicological studies are in
Table 8-3.
8.3.5.1 Epidemiologic Studies on Spontaneous Abortion and Pregnancy Loss and
Fetal and Infant Mortality
In the 2013 Pb ISA, there was a limited number of epidemiologic studies that examined Pb
exposure and spontaneous abortion or pregnancy loss with inconsistent findings. Studies that examine
spontaneous abortion or pregnancy loss are difficult to conduct, as many spontaneous abortions or
pregnancy losses occur during the first trimester. Women may miscarry before being enrolled in a study
and/or women may not have known they were pregnant when they miscarried, further limiting the ability
to detect subtle effects, especially if higher Pb exposures do lead to increased risk of early spontaneous
abortions or pregnancy loss. In addition, some studies are limited by their retrospective examination of
current Pb biomarker levels in relation to previous miscarriages. The epidemiologic studies reviewed in
the 2013 Pb ISA had limited sample sizes and little control for potential confounding factors, with some
studies including no potential confounders in their analyses. There were no epidemiologic studies of Pb
exposure and fetal and infant mortality reviewed in the 2013 Pb ISA and there were no recent PECOS-
relevant epidemiologic studies of Pb exposure and fetal and infant mortality.
There were only a few recent epidemiologic studies that evaluated Pb exposure and spontaneous
abortion and pregnancy loss. There were inconsistent findings among the studies and no apparent pattern
of association by biomarker of Pb exposure. In a recent study, cord blood samples were obtained from
432 infants born in an area with e-recycling (Guiyu) and 99 from an area without e-recycling (Xiamen) in
China (Xu et al.. 2012). There was an increased risk of 4.20 (95% CI: 3.40, 5.18) of stillbirth rate with
cord BLLs comparing infants from the area with e-recycling (Guiyu) compared to infants from the area
without e-recycling (Xiamen). In a recent cohort study, couples (n = 344) were prospectively followed to
explore the relationship between blood Pb concentrations at enrollment and with pregnancy followed to
estimate the risk of incident of pregnancy loss (Louis et al.. 2017). Each participant's blood Pb
concentration and time to pregnancy loss was modeled individually and as a couple. In the individual
partner models, there was no increased risk of pregnancy loss for female partner blood Pb (hazard ratio
[HR]: 1.01 [95% CI: 0.82, 1.25]) or male partner blood Pb (HR: 0.95 [95% CI: 0.77, 1.17]). In the
couple-based model, the associations were unchanged (female partner HR: 1.01 [95% CI: 0.80, 1.28] and
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male partner HR: 0.96 [95% CI: 0.77, 1.22]). Among a cohort of 166 women in Iran, there was no
increased risk (OR: 1.08 [95% CI: 0.98, 1.20]) of spontaneous abortion with maternal BLLs in early
pregnancy (Vigeh et al.. 2021). In another prospective cohort among women seeking treatment at a
fertility clinic in Turkey, blood Pb concentrations were assessed in association with ongoing pregnancy
(Tolunav et al.. 2016). The study participants were categorized into patients with ongoing pregnancy
(n = 20) and patients who experienced assisted reproductive technology (ART) failure, miscarriage, or
biochemical pregnancy (n = 81). There was a 2.2% lower risk (relative risk [RR]: 0.978 [95% CI: 0.957,
0.999]) for ongoing pregnancy for each 1 (ig/dL higher blood Pb concentration. Among a cohort of 1,184
women undergoing assisted reproductive therapy in China, associations between maternal serum Pb
concentrations and spontaneous abortion before gestational week 12 were evaluated (Li et al.. 2022).
There was an increased risk of 1.39 (95% CI: 1.02, 1.91) of spontaneous abortion before gestational week
12 with increasing maternal Pb serum levels. When categorized into tertiles, the associations between
maternal Pb serum levels and spontaneous abortion before gestational week 12 were null.
8.3.5.2 Toxicological Studies on Spontaneous Abortion and Pregnancy Loss and
Fetal and Infant Mortality
The 2013 Pb ISA did not report any toxicological studies that investigated the effects of Pb on
offspring mortality at any stage of development. Some studies that investigated the effects of Pb on
offspring mortality were summarized in the 2006 Pb AQCD. Overall, these studies found that gestational
exposure increased pregnancy loss and implantation loss (BLLs >32 (ig/dL) (Pinon-Lataillade et al..
1995; Singh et al.. 1993; Piasek and Kostial. 1991; Logdberg et al.. 1987). Some recent studies have also
investigated the effects of Pb exposure on offspring mortality (Table 8-3). However, recent studies
reported that Pb did not have effects on measures of pre- or postnatal mortality, including litter size.
Rodent studies that dosed prior to and during gestation reported no increase in stillbirth or decrease in
number of pups born to treated dams (Saleh et al.. 2018; Rao Barkur and Bairv. 2016; Barkur and Bairv.
2015; Weston et al.. 2014; Corv-Slechta et al.. 2013; Betharia and Maher. 2012). BLLs, sources
(e.g., BLLs from dams or BLLs from offspring), and times of measurement were variable between these
studies (0.0318-27.7 (j,g/dL; GD 20-PND 60), but in general, BLLs in recent studies were lower than
those reported in previous studies. The contrast in the effects of Pb exposure on offspring mortality
observed between previous studies and recent studies may be attributed to the lower BLLs achieved in
recent studies compared to the higher BLLs in previous studies.
Postnatal offspring mortality was also investigated in some rodent studies, and some studies
reported on measures of offspring mortality that included postnatal death and survival until certain
timepoints after birth (e.g., weaning). These studies also did not report any effects of Pb exposure on
postnatal survival. Most studies utilized dosing paradigms that dosed before or during gestation (Barkur
and Bairv. 2015; Betharia and Maher. 2012) and reported BLLs at different times postnatally (PND 2-60;
0.0318 (ig/dL-5.30 (ig/dL) with BLLs tending to be lower in time points with the longest amount of time
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since cessation of exposure. Some studies utilized a dosing paradigm that exclusively exposed animals
postnatally (Barkur and Bairv. 2015; Graham et al.. 2011). In agreement, these studies also reported no
effects of offspring mortality during postnatal time points (PND 4-29; BLLs 3.27-26.65 (.ig/dL).
8.3.5.3 Integrated Summary of Effects on Spontaneous Abortion and Pregnancy Loss
and Fetal and Infant Mortality
The 2013 Pb ISA reported inconsistent findings from the epidemiologic studies on Pb exposure
and spontaneous abortion and pregnancy loss. The findings from recent epidemiologic studies on Pb
exposure and spontaneous abortion and pregnancy loss were also inconsistent. A single study reported
increased risk of stillbirth with cord blood Pb. While recent cohort studies among healthy participants did
not find increased risk of pregnancy loss or spontaneous abortion, women seeking treatment from a
fertility clinic had increased risk of spontaneous abortion before gestational week 12 or decreased risk of
an on-going pregnancy. The women seeking fertility treatment that were recruited as participants may be
different from those in the general population, limiting the generalizability of the results as the study
populations may not be representative of the general population as they have already been diagnosed and
are seeking treatment for fertility issues. However, early pregnancy loss is more likely to be ascertained
from women seeking treatment at fertility clinics. Previous toxicological studies reported increased rates
of pregnancy loss and implantation loss in animals dosed with Pb during gestation. This contrasts with
more recent literature which did not report any effect of Pb on pre- or postnatal offspring mortality.
Although not always consistent, BLLs were generally lower in recent toxicological literature when
compared to previous literature, possibly explaining the observed contrast in results.
8.3.6 Placental Function
In the 2013 Pb ISA, there were no epidemiologic or toxicological studies available that evaluated
Pb concentrations and associations with placenta function. Recent epidemiologic and toxicological studies
evaluating the association between Pb exposure and placental function are limited. The epidemiologic
studies were cross-sectional studies. Study details for the recent epidemiologic studies are included in
Table 8-8 and the toxicological studies are included in Table 8-3.
8.3.6.1 Epidemiologic Studies on Placental Function
In the 2013 Pb ISA, there were no epidemiologic studies available that evaluated Pb
concentrations and associations with placenta function. In recent cross-sectional epidemiologic studies,
there were different markers of placental function evaluated. One marker of placental function that was
evaluated was placental thickness, which can restrict intrauterine fetal growth (Al-Saleh et al.. 2014).
Maternal BLLs measured at delivery were found to be associated with the risk of placental thickness
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below the 10th percentile (OR: 1.64 [95% CI: 1.12, 2.41]). In another study, using a cross-section from
the JECS, the relationship between maternal blood Pb collected during the second trimester and placental
previa and placenta accreta among 16,019 women was examined (Tsuii et al.. 2019). Placenta previa is a
condition in which the placenta is attached to the lower uterine segment and completely or partially
covers the internal cervix, and when chorionic villi abnormally invade to myometrium, placenta accreta
occurs (Tsuii et al.. 2019). There was increased odds of placenta previa in the second quartile (4.80-
5.95 ng/g) of maternal blood Pb (OR: 2.59 [95% CI: 1.40, 4.80]), but null associations the third quartile
(5.96-7.44 ng/g) maternal blood Pb (OR: 1.32 [95% CI: 0.66, 2.64]) and fourth quartile (>7.45 ng/g)
maternal blood Pb (OR: 1.34 [95% CI: 0.67, 2.67]). There were null associations for placenta accreta
across the blood Pb quartiles (Table 8-8).
8.3.6.2 Toxicological Studies on Placental Function
The 2013 Pb ISA reported a single study that investigated the effects of Pb exposure on placental
function. Wang et al. (2009) reported decreased placental weight in Wistar rats along with dose-
dependent increasing pathology of cytoarchitecture and cytoplasmic organelles. Focusing on different
gestational periods, this study exposed dams to 0.025% Pb via drinking water from either GD 1-10,
GD 11-20, or GD 1-20 (maternal BLLs on GD 20 were 26.3, 12.4 (ig/dL, and 36.0 (ig/dL, respectively).
Some recent studies have reported on similar placental outcomes (Table 8-3). Wang et al. (2014) also
dosed using the same dosing paradigm (dosing during GD 1-10, GD 11-20, or GD 1-20 via drinking
water) and reported that placentae collected from pregnant Wistar rats on GD 20 showed similar dose-
dependent decreases in weight and histopathological abnormalities such as vascular congestion,
trophoblast degeneration, chorionic villi interstitial edema, irregularity of trophoblast cells in the labyrinth
and trophospongium, degeneration of trophoblast cells, and chorionic villi vacuolization (maternal Pb
levels on GD 20 were reported to be between 12.4-36.0 (ig/dL and varied by dosing window). Two other
studies that dosed pregnant Sprague-Dawley rats via gavage from GD 0-20 and similarly reported
reduced placental weights (maternal blood Pb on GD 20 was 23.9-27.7 (ig/dL) (Saleh et al.. 2019; Saleh
et al.. 2018). Of note is that both of these recent studies by Saleh et al. (Saleh et al.. 2019; Saleh et al..
2018) also reported reduced brain weights in dams which is indicative of overt toxicity. Thus, it is
possible the altered placental weight could be attributed to overt toxicity experienced by the dams.
8.3.6.3 Integrated Summary of Effects on Placental Function
There were no epidemiologic studies available that evaluated Pb concentrations and associations
with placenta function in the 2013 Pb ISA. The recent epidemiologic studies reviewed that assessed the
relationship of Pb exposure and placental are limited. The differences in the different markers of placental
function make it difficult to judge coherence and consistency across these studies, but these positive
associations are an indication that exposure to Pb may result in effects on placental function during
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pregnancy. Previous toxicological data on the effects of Pb on placental weight are limited to a single
study which reported decreased placental weight and histological alterations. Recent studies also reported
that dams dosed with Pb had reduced placental weight, but of note is that these studies also reported
reduced brain weight in dams, suggesting that overt toxicity may have occurred and could be related to
the observed reductions in placental weight.
8.3.7 Other Pregnancy and Birth Outcomes
There were several recent studies that evaluated associations between Pb exposure and other
pregnancy and birth outcomes in the epidemiologic and toxicological literature. More specific study
details for the epidemiologic studies, including Pb levels, study population characteristics, potential
confounders, and select results from these studies are highlighted in Table 8-9. Specific study details for
the toxicological studies are provided in Table 8-3.
8.3.7.1 Epidemiologic Studies on Other Pregnancy and Birth Outcomes
There were several recent studies with other outcomes related to pregnancy and birth. More
specific study details, including Pb levels, study population characteristics, potential confounders, and
select results from these studies are highlighted in Table 8-9. In studies of other pregnancy and birth
outcomes, maternal Pb blood concentrations were associated with high levels of leptin, a fetal marker of
metabolic function (Ashley-Martin et al.. 2015a); and cord blood Pb concentrations were negatively
associated with cord blood relative telomere length (rTL) (Herlin et al.. 2019). However, there were null
associations in several other studies evaluating Pb exposure and outcomes related to pregnancy and birth.
There was a null association between maternal serum Pb levels and nuchal translucency, which is the
subcutaneous space in the fetal neck and is visible with ultrasound imaging in the first trimester (Liaoet
al.. 2015). Increased nuchal translucency thickness in the first trimester has been reported to be a risk
factor for chromosomal abnormalities, genetic syndromes, congenital heart defects, structural
abnormalities, intrauterine infection, neurodevelopmental delay, and fetal demise (Liao et al.. 2015).
There was null associations between maternal blood Pb concentrations and thymic stromal lymphopoietin
(TSLP) and interleukin-33 (IL-33), which are biomarkers of fetal immune system (Ashley-Martin et al..
2015b); maternal blood Pb concentrations and elevated cord blood concentrations of immunoglobulin E
(IgE) (Ashley-Martin et al.. 2015b); and maternal blood Pb and markers of fetal metabolic function (low
leptin, low adiponectin, and high adiponectin) (Ashley-Martin et al.. 2015b). Additionally, there were
inconsistent associations between maternal BLLs during pregnancy and secondary sex ratio. Among
participants in the ALSPAC in the United Kingdom, Taylor et al. (2014) reported no associations among
quintiles of maternal Pb levels during the first trimester of pregnancy and secondary sex ratio (odds of
having a male child). Among the participants of the Longitudinal Investigation of Fertility and the
Environment (LIFE) cohort, there were associations with secondary sex ratio (ratio of live male to female
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births, reflecting a male excess) and both maternal and parental blood samples were measured for Pb at
baseline (before pregnancy) (Bloom et al.. 2015). However, Tatsuta et al. (2022b) reported increased odds
of male births (secondary sex ratio) of 1.279 (95% CI: 1.224, 1.336) in the highest quintile of maternal
blood Pb among a subset of participants in the JECS.
8.3.7.2 Toxicological Studies on Other Pregnancy and Birth Outcomes
The 2013 Pb ISA summarized a single study reporting on other birth outcomes such as sex ratio.
Dumitrescu et al. (2008a) reported an increased female:male ratio in offspring born to Wistar rats that
were dosed with 100 or 150 ppb of Pb in the drinking water starting 3 months prior to mating and
continuing until birth. This study did not report on BLLs in dams or offspring, resulting in more
challenges when comparing this study to more recently published studies (Table 8-3). Weston et al.
(2014) was the only recent study to report an effect of Pb exposure on the sex ratio of pups born to
exposed dams. In agreement with Dumitrescu et al. (2008a). Weston et al. (2014) reported that Long-
Evans rats dosed starting 76 days prior to mating and continuing through birth gave birth to female
skewed litters when compared to control. However, it is worth noting that in Weston et al. (2014) control
litters had unusually high numbers of males that resulted in a 1.5 male:female ratio of pups born, whereas
Pb-treated litters had a more even distribution that resulted in a ratio closer to 1:1 (BLLs were 14.6 and
15.7 (ig/dL in female and male pups, respectively, on PNDs 5-6). Other recent studies contrast with these
studies and report no effects on sex ratio in Pb-treated females. One study using a dosing paradigm
similar to the two studies above (exposure beginning 2 months prior to dosing and continuing through
birth) reported no effects of Pb on sex ratio (BLLs 12.12 (ig/dL in dams at weaning) in C57BL/6 mice
(Corv-Slechta et al.. 2013). Similarly, Tartaglione et al. (2020) dosed Wistar rats from 4 weeks prior to
mating through birth and reported no changes in sex ratio (BLLs 25.5 (ig/dL on PND 23 in pups).
Additional studies in rats dosed females from the beginning of pregnancy through birth and also reported
no changes in sex ratio (BLLs 6.68-9.03 (ig/dL in pups taken at ages PND 2 and PND 28 in (Betharia and
Maher. 2012) and (Baranowska-Bosiacka et al.. 2013). respectively). With the only recent study reporting
alterations in sex ratio also reporting unusual sex ratios in the control group, the effects of Pb exposure on
sex ratio is equivocal.
8.3.7.3 Integrated Summary of Effects on Other Pregnancy and Birth Outcomes
There was a small body of recent epidemiologic studies across various other pregnancy and birth
outcomes; however, the small number of studies limits the ability to judge coherence and consistency
across these studies, although the associations reported demonstrate that Pb exposure could result in
physiological responses that contribute to adverse pregnancy and birth outcomes, such as markers of fetal
metabolic function, fetal immune system biomarkers, and rTL. Toxicological evidence regarding other
pregnancy and birth outcomes are equivocal. While the 2013 Pb ISA reported a study that found that Pb
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exposure led to female-skewed litters, a few recent studies reported no effects of Pb on the ratio of male
to female pups born to Pb-exposed dams.
8.4 Effects on Development
The 2013 Pb ISA (U.S. EPA. 2013) concluded that the collective body of evidence integrated
across epidemiologic and toxicological studies, based on the findings of delayed pubertal onset among
males and females, was sufficient to conclude that there is a causal relationship between Pb exposure and
developmental effects. The current epidemiologic and toxicological studies continue to support
associations between Pb exposure and developmental effects, particularly the delayed onset of puberty in
both males and females. This section does not cover associations between Pb exposure and
neurodevelopmental outcomes, which are discussed in detail in Appendix 3 Nervous System Effects. The
recent epidemiologic and toxicological studies of Pb exposure and effects on development are detailed in
the following sections. The developmental endpoints in subsequent sections are based on postnatal
growth, bodyweight, and stature; puberty among females; and puberty among males.
8.4.1 Effects on Postnatal Growth
The recent epidemiologic and toxicological studies that examine the relationship between Pb
exposure and postnatal growth are detailed below. More specific study details for the epidemiologic
studies, including Pb levels, study population characteristics, potential confounders, and select results
from these studies are highlighted in Table 8-10. Specific study details for the toxicological studies are
provided in Table 8-11.
8.4.1.1 Epidemiologic Studies on Postnatal Growth
The 2013 Pb ISA found inconsistent results between Pb exposure and postnatal growth.
Longitudinal epidemiologic studies had inconsistent findings regarding the association between Pb levels
and postnatal growth. There were further inconsistencies in the findings of the cross-sectional studies
evaluated. While multiple cross-sectional studies reported an association between Pb levels and impaired
growth, several other cross-sectional studies did not report associations between Pb and growth. The
inconsistencies across the studies may be due to study design and differences in the timing of exposure to
Pb (e.g., prenatal, at delivery, or postnatal). However, the longitudinal studies were controlled for
multiple potential confounders, such as age and parity.
There were multiple recent epidemiologic studies that evaluated the relationship between Pb
exposure and postnatal growth. Overall, there were negative associations between Pb exposure and
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specific postnatal growth outcomes among the cross-sectional studies. However, among cohort studies,
there were some inconsistencies in the associations of Pb exposure and different postnatal growth
outcomes. These inconsistencies in the cohort studies may be due to differences in the timing of when Pb
exposure was measured, the biomarker of Pb exposure, and the timing of the outcome.
Among the other cross-sectional studies of Pb exposure and postnatal growth, there were
consistent negative associations. In a National Health and Nutrition Examination Survey (NHANES)
(2013-2016) analysis of 6-11-year-old children (n = 1,634), there were negative associations between an
IQR difference in blood Pb concentrations (median: 0.5 (ig/dL) and standing height (|3: -3.116 cm [95%
CI: -5.03, -1.202]), waist circumference (WC; |3: -5.742 cm [95% CI: -8.769, -2.715]), upper arm
length (|3: -1.068 cm [95% CI: -0.625, -0.512]), and BMI (|3: -2.092 kg/m2 [95% CI: -3.227, -0.957])
(Signes-Pastor et al.. 2021). Among male participants, the negative associations persisted with postnatal
growth outcomes. Among female participants, there were negative associations with BMI, WC, and upper
arm length, but null associations with standing height. Similarly, in a study of primary school children
aged 7-11 years in China, there were negative associations with concurrently measured BLLs (median:
2.61 ng/dL) and height (|3: -3.21 cm [95% CI: -4.24, -2.17]), weight (|3: -1.96 kg [95% CI: -3.11,
0.82]), bust circumference (|3: -2.77 cm [95% CI: -3.79, -1.76]), and waistline (|3: -3.65 cm [95% CI:
-4.78, 2.52]); however, there was a null associations with BMI (|3: 0.20 kg/m2 [95% CI: -0.65, 0.25])
(Kuang et al.. 2020).
When standardizing postnatal growth metrics by Z-score, the associations with Pb exposure were
mixed, even with all median BLLs across studies less than 5 (ig/dL (range: 0.663-4.6 (ig/dL). In a cross-
sectional study, among children <6 years of age (n = 1,678) in China, there were negative associations
between children's logio-BLLs and weight for age Z-score (WAZ) (|3: -0.33 [95% CI: -0.56, -0.11]) and
height-for-age Z-score (HAZ) (|3: -0.38 [95% CI: -0.63, -0.14]), but null associations with BMI-for-age
Z-score (BMIZ) (Zhou et al.. 2020). When the BLLs were grouped by tertiles, the children in the highest
tertile (>5 ^ig/dL) had lower WAZ (|3: -0.42 [95% CI: -0.62, -0.23]), lower HAZ (|3: -0.36 [95% CI:
-0.58, -0.15]), and lower BMIZ (|3: -0.29 [95% CI: -0.50, -0.07]) than those in the lowest tertile
(<2.5 (ig/dL). The patterns of association held when stratified by child's sex (see Table 8-10). Among
children ranging in age from 8 to 23 months in South Korea, BLLs were associated with post-birth weight
gain (WAZ-BWZ, or the difference of the WAZ at the time of the study and BWZs) (|3: -0.238 [-0.391,
-0.085], standard error [SE]: 0.078) and current HC for age Z-scores (HCAZs; |3: -0.213 [-0.366, -0.06],
SE:0.078) (Choi et al.. 2017). However, among participants in the Canadian MIREC Child Development
Plus Study, there were no associations reported between blood Pb measured at 2 and 5 years of age and
HAZ, WAZ, or BMIZ overall or when stratified by child's sex (Ashley-Martin et al.. 2019).
Multiple cohort studies examined the relationship between Pb exposure at different time periods
(prenatally and/or at different time periods during childhood) with growth metrics, mainly height and
weight. Among the cohort studies that measured Pb in cord blood, there were inconsistent associations.
While a study among children in Krakow, Poland found no associations with change in mean height over
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a 9-year follow-up period (Jcdrvchowski et al.. 2015). a study in the Children's Health and Environmental
Chemicals in Korea (CHECK) study, reported positive associations with Z-scores for weight and BMI at
24 months of age (weight |3: 0.717 [95% CI: 0.195, 1.239] and BMI |3: 0.695 [95% CI: 0.077, 1.313],
respectively) (Kim et al.. 2017). However, there were no associations between cord blood Pb and the Z-
scores of the child's weight, height, or BMI at any other time point (see Table 8-10 for details). When
stratified by children's sex, cord blood Pb was positively associated with an increase of birth height (|3:
0.017 [95% CI: 0.003, 0.031]) and a decrease of PI at birth (|3: -0.055 [95% CI: -0.103,-0.006]) in boys,
but not in girls.
Among participants (n = 1,150) in the Mothers' and Children's Environmental Health (MOCEH)
study in South Korea, maternal BLLs at delivery were negatively associated with Z-scores of weight for
age (|3: -0.33 [95% CI: -0.53, -0.13]) and length for age (|3: -0.30 [95% CI -0.53, -0.08]) at 24 months,
meaning that a l-(ig/dL increase in late pregnancy Pb levels decreased weight and length at 24 months by
0.33 kg and 0.30 cm, respectively (Hong et al.. 2014). However, there were no associations between
maternal BLLs in early pregnancy (before 20 weeks gestation) and cord blood Pb with weight and length
(Table 8-10).
Several cohort studies conducted in Mexico measured Pb exposure during pregnancy in maternal
blood, cord blood, and maternal bone and various postnatal growth outcomes. Renzetti et al. (2017)
investigated how Pb exposure during pregnancy is associated with children's growth outcomes, including
height, weight, BMI, and percentage body fat, measured between ages 4-6 years old in a Mexico City
pregnancy cohort (PROGRESS). Maternal blood Pb was measured during the second and third trimester
of pregnancy, as well as at delivery. Cord blood was measured at delivery. Bone Pb levels in the tibia and
patella were also assessed in mothers as a long-term biomarker 1 month postpartum. There were negative
associations between maternal third trimester BLLs and height-for-age (|3: -0.10 [95% CI: -0.19, -0.01])
and weight for age (|3: -0.11 [95% CI: -0.22, -0.003]), but there were no associations between any other
marker of Pb exposure (maternal second trimester blood, cord blood Pb, maternal blood at delivery) and
height-for-age, weight for age, BMI, or percentage of body fat (see Table 8-10). In the Early Life
Exposure in Mexico to Environmental Toxicants (ELEMENT) project, Liu et al. (2019a) assessed Pb
exposure in maternal bone (as a proxy for cumulative fetal exposure) at 1 month postpartum and also in
blood samples from children annually from 1 to 4 years in association with BMIZ, WC, sum of skinfolds,
and body fat percentage in 248 children aged 8-16 years. Maternal patella Pb levels were associated with
lower child BMIZ (|3: -0.02 [95% CI: -0.03, -0.01]), WC (|3: -0.12 cm [95% CI: -0.22, -0.03]), sum of
skinfolds (|3: -0.29 mm [95% CI: -0.50, -0.08]), and body fat percentage (|3: -0.09% [95% CI: -0.17,
-0.01]). However, there were no associations detected from the postnatal exposure period (blood samples
in children). In another study, children born between 1994 and 2005 in Mexico City had Pb exposure
measured in maternal patella Pb concentrations, a marker of prenatal period exposure, and from infant
and childhood measured in blood at birth to 24 months and 30-48 months (Afeiche et al.. 2012). Among
infants with BLL exceeding the median (4.5 (.ig/dL). there was a decrease in height of 0.84 cm (95% CI:
-1.43, -0.26) compared to children with a level below the median. There were no associations between
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prenatal Pb or childhood Pb and height and there were no associations with BMI at any time point
(prenatal, infancy, or childhood). In cohort of Mexican children aged 6-8 years old, growth (height, HAZ,
and knee height) were assessed in association with BLLs at baseline, after 6 months, and 12 months (Kerr
et al.. 2019). Additionally, as BLLs may differ by the aminolevulinic acid dehydratase (ALAD) genotype,
the authors compared children with the ALAD 1-2/2-2 genotype to children with the ALADi 1 genotype.
There were negative associations with height (|3: -0.11 cm [95% CI: -0.18, -0.04]), knee height (|3:
-0.04 cm [95% CI: -0.07, -0.02]), and HAZ (|3: -0.02 cm [95% CI: -0.03, -0.01]). Children with
ALADi 1 had decreased height, knee height, and HAZ, while children with the A LA D1 2 2 2 had reduced
knee height and HAZ, but not height. There were no associations between BLLs and growth at 6- or 12-
month follow-up reported, irrespective of ALAD genotype. This epigenetic study proposes a potential
mechanistic pathway of BLLs differing by genotypes and the associations with growth metrics during
child developmental periods.
There were several studies that explored the associations between Pb exposure and postnatal
growth specifically by sex. In a study by Burns et al. (2017). associations of BLLs were assessed with
longitudinal age-adjusted height (HAZ) and BMI (BMIZ) among male participants in the Russian
Children's Study. Over 10 years of follow-up, after covariate adjustment, boys with higher (>5 (ig/dL)
BLLs compared with lower BLLs were shorter (adjusted mean difference in HAZ: -0.43 [95% CI -0.60,
-0.25]), translating to a 2.5 cm lower height at age 18 years. The decrement in height for boys with higher
BLLs was most pronounced at 12 to 15 years of age (interaction p: 0.03). However, boys with higher
BLLs were leaner (adjusted mean difference in BMIZ: -0.22 [95% CI: -0.45, 0.01]). Deierlein et al.
(2019) used data from the Breast Cancer and the Environment Research Program to investigate
associations of childhood blood Pb concentrations and anthropometric measurements among a multi-site,
multiethnic cohort of girls (n = 683). Blood Pb concentrations were collected before 10 years of age and
height, BMI, WC, percent body fat was measured between 7-14 years of age. There were decreases in
height (range: -2.0 to -1.5 cm), BMI (range: -0.9 to -0.7 kg/m2), WC (range: -3.0 to -2.2 cm), percent
body fat (range: -2.9 to -1.7%) among girls ages 7 through 14 with BLLs of >1 (ig/dL compared to
<1 |ig/dL (Table 8-10).
There were a limited number of studies that examined stunting with exposure to Pb in children. A
single cross-sectional study in a subset of participants in the Interactions of Malnutrition and Enteric
Infections: Consequences for Child Health and Development (MAL-ED) study in Bangladesh reported
increased odds of stunting (OR: 1.78 [95% CI: 1.07, 2.99]) and being underweight (OR: 1.63 [95% CI:
1.02, 2.61]) with elevated (>5 (ig/dL) BLLs, but not wasting (OR: 1.18 [95% CI: 0.64, 2.19]) (Raihan et
al.. 2018). In a cohort study among rural Bangladeshi children, Pb exposure was assessed from umbilical
cord blood at birth and blood Pb at 20-40 months of age with stunting (Gleason et al.. 2016). The odds of
stunting at 20-40 months was 1.12 (95% CI: 1.02, 1.22) per each 1 (ig/dL increase in childhood BLL;
however, there was no association was found between cord BLL and risk of stunting (OR: 0.97 [95% CI:
0.94-1.00]).
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8.4.1.2
Toxicological Studies on Postnatal Growth
The 2013 Pb ISA summarized several current studies and several from the 2006 Pb AQCD that
reported on the effects of Pb on offspring bodyweight and size. The reported effects were fairly
consistent, and nearly all studies reported reductions in bodyweight of offspring exposed to Pb during
developmental periods. Gestational exposure to Pb proved to be sufficient to reduce offspring bodyweight
and body length in rodent studies (Masso-Gonzalez and Antonio-Garcia. 2009; Wang et al.. 2009; Teiion
et al.. 2006; Ronis et al.. 2001; Ronis et al.. 1998a; Ronis et al.. 1998c; Ronis et al.. 1996). The only study
in the 2013 Pb ISA that did not report a reduction in bodyweight was Leasure et al. (2008). which
reported that bodyweight increased at 1 year of age in male C57BL/6 offspring that were dosed starting
prior to conception and ending on PND 10.
Many recent publications have also reported on the effects of Pb-induced changes in bodyweight
of offspring (Table 8-11). In contrast with the 2013 Pb ISA, only some studies reported that Pb exposure
affected offspring bodyweight. One study that dosed Sprague-Dawley rat pups with 1 or 10 mg/kg/d
directly via gavage from PND 4 to 28 reported that bodyweight was decreased in male offspring in both
groups on PND 26 (BLLs 3.27-12.5 (ig/dL on PND 29) (Graham et al.. 2011). Another study dosed
Wistar rat dams via drinking water (30 mg/L Pb) from birth until weaning, at which point offspring were
weaned onto the same dosage of Pb in their drinking water as their dam until outcome assessment (de
Figueiredo et al.. 2014). In agreement with Graham et al. (2011). this study by de Figueiredo et al. (2014)
reported reductions in bodyweight of male Wistar rat offspring on PND 60 that were exposed from
conception through PND 60, although female offspring were not evaluated (BLLs 7.2 (ig/dL on PND 60).
Similarly, one study exposed CD-I mice offspring to Pb via dam drinking water (27 or 109 ppm Pb) from
PND 1 to 21 and reported reductions in body weight of pups on PNDs 11, 15, and 19 (BLLs 19.57—
29.16 (ig /dL on PND 18) (Duan et al.. 2017). In contrast, one study conducted in Sprague-Dawley rats
that were exposed from GD 0 to PND 21 via the dam's drinking water (10 |ig/m L Pb) reported increased
bodyweight in offspring on PND 1 and increased bodyweights in females only on PND 49 and 56 (BLLs
9.03 (ig/dL on PND 2, 0.976 (ig/dL on PND 25, 0.0318 (ig/dL on PND 60 in pups) (Bctharia and Mahcr.
2012).
Contrasting these recent studies and previous studies discussed in the 2013 Pb ISA are several
studies that reported no effects of Pb exposure on bodyweight in offspring. Studies in both mice and rats
utilizing dosing paradigms that begin exposure prior to conception (Albores-Garcia et al.. 2021; Zhao et
al.. 2021; Sobolewski et al.. 2020; Rao Barkur and Bairv. 2016; Barkur and Bairv. 2015) reported no
effects of Pb exposure on bodyweight at any time point measured (BLLs ranged between 0.4-15.7 (ig/dL
across various time points and studies). Similar null findings were reported in other rodent studies
utilizing other exposure windows including gestational (Rao Barkur and Bairv. 2016; Barkur and Bairv.
2015). lactation (Rao Barkur and Bairv. 2016; Barkur and Bairv. 2015; Basgen and Sobin. 2014). and a
combination thereof (Basha and Reddv. 2015; Barkur etal.. 2011) (BLLs ranged between 2.74-
26.86 ng/dL).
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8.4.1.3 Integrated Summary of Effects on Postnatal Growth
Overall, among the recent epidemiologic studies, there were negative associations between Pb
exposure and specific postnatal growth outcomes among the cross-sectional studies. However, among
cohort studies, there were some inconsistencies in the associations of Pb exposure and different postnatal
growth outcomes. These inconsistencies in the cohort studies may be the result of the differences in the
timing of Pb exposure measurement (prenatally or postnatally) and the biomarker to measure Pb exposure
(maternal blood, maternal bone, cord blood). Additionally, there was limited evidence that there are
potential differences in the associations between Pb exposure and growth metrics between males and
females. There is also limited evidence of potential epigenetic effects of BLLs differing by genotypes and
the associations with growth metrics during child developmental periods. While cross-sectional studies
are limited by the concurrent measurement of Pb and postnatal growth outcomes, there were several well-
designed cohort studies that support the associations of Pb exposure and decreased growth. These studies
accounted for a wide range of potential confounders, including co-exposure to other metals; however,
some studies did not consider prenatal growth (birth weight, birth length) or maternal characteristics
(height, weight, BMI, smoking), which could potentially influence postnatal growth. While there was a
small body of literature examining the associations between stunting and exposure to Pb, there was
consistent increased odds in stunting with Pb exposure. Previous toxicological studies tended to report
reductions in postnatal weight of offspring exposed to Pb; however, recent literature is inconsistent. Some
studies reported reductions of offspring weight following exposure to Pb in prenatal or early postnatal
life, while others report no effects of Pb on postnatal weight in offspring. Discerning reasons for the
observed inconsistencies is difficult because studies still reported results that contrasted with other studies
that used similar dosing windows, doses, and animal species.
8.4.2 Effects on Puberty among Females
The recent epidemiologic and toxicological studies examining the relationship between Pb
exposure and effects on puberty among females are summarized in the text below. Study details of the
recent epidemiologic studies are included in Table 8-12.
8.4.2.1 Epidemiologic Studies on Puberty among Females
The epidemiologic studies reviewed in the 2013 Pb ISA found consistent associations between
higher concurrent blood Pb and delayed pubertal development in females. The association persisted in
populations with mean and/or median concurrent BLLs of 1.2-9.5 (ig/dL. While most of the studies had
large sample sizes and controlled for potential confounders, they were cross-sectional study designs, so
there are some uncertainties regarding temporality between Pb exposure and pubertal onset; additionally,
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these studies were not able to separate out the influence of past Pb exposure, including prenatal
exposures, from more recent exposures.
The recent epidemiologic studies assessing the associations between blood Pb and onset of
puberty among females used different markers of puberty. A single cross-sectional study of NHANES
(2011-2012) data evaluated the associations between blood Pb concentration and circulating serum total
testosterone levels in 6-19-year-old children and adolescents (Yao et al.. 2019). Testosterone is a
principal sex hormone needed for normal physiologic processes during all lifestages and for females,
testosterone is of crucial importance for bone density and necessary for normal ovarian and sexual
function, libido, energy, and cardiovascular and cognitive functions (Yao et al.. 2019). While there were
no associations between blood Pb and testosterone in female children (6-11 years), serum testosterone
levels were 14.85% greater (95% CI: 0.83%, 30.81%) in female adolescents (12-19 years) in the lowest
quartile of BLLs (<0.35 (ig/dL) than those in the highest quartile (>0.63 (ig/dL) in the fully adjusted
model. For both female children and adolescents, there were no significant trends with increasing
quartiles of exposure (p for trend 0.63 and 0.08, respectively).
In a cross-sectional study in Poland, two different groups of adolescent girls aged 7-16 years
(n = 436 in 1995 and n = 361 in 2007) were assessed for effects of Pb on the age at attaining menarche
(Slawinska et al.. 2012). While the associations between blood Pb and menarche from either group of
girls were null (1995 OR: 0.70 [95% CI: 0.27, 1.85] and 2007 OR: 0.31 [95% CI: 0.09, 1.06]), the
patterns of association between these two times periods suggest delayed menarche. In another cross-
sectional study among school-age girls (n = 490) in Poland, there was a pattern of decreased odds
between blood Pb and age at menarche, whether controlled for BMI (OR: 0.54 [95% CI: 0.26, 1.13]),
percentage of body fat (OR: 0.52 [95% CI: 0.25, 1.08]), or sum of skinfolds (OR: 0.53 [95% CI: 0.26,
1.10]) (Gomula et al.. 2022). While these cross-sectional studies reported imprecise associations, the
pattern of association is important to note.
Three successive, cross-sectional Flemish Environment and Health Studies (FLEHS I, FLEHS II,
and FLEHS III) were conducted among adolescents (aged 14-15 years old) in Belgium between 2002-
2015 (De Craemer et al.. 2017). Female puberty markers of age at menarche, breast development, and
pubic hair development were evaluated in relation to blood Pb exposure. There was a consistent pattern of
delayed age at menarche across the three study cohorts (FLEHS I OR: 0.039 [95% CI: -0.072, 0.15];
FLEHS II OR: 0.257 [95% CI: 0.091, 0.424]; FLEHS III OR: 0.126 [95% CI: -0.021, 0.273]). The
associations between blood Pb and breast development were inconsistent, but there was indication of
delayed development among FLEHS I participants (OR: 0.798 [95% CI: 0.653, 0.969]). There were no
associations between blood Pb and development of pubic hair among adolescent females across the three
study cohorts.
Multiple cohort studies examined the associations between Pb exposure and puberty in females.
These studies used different biomarkers of exposure and different markers of puberty (Liu et al.. 2019b:
Jansen et al.. 2018: Nkomo et al.. 2018). Cord BLLs and BLLs at age 13 were evaluated in association
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with puberty progression (development of pubic hair and development of breasts) among 684 females in
the Birth to Twenty Plus (BT20+) birth cohort in South Africa (Nkomo et al.. 2018). In females with
elevated BLLs (>5 (ig/dL) at age 13, there was lower level of breast development (RR: 0.45 [95% CI:
0.29, 0.68]) and slower progression of pubic hair development (RR: 0.46 [95% CI: 0.27, 0.77]), but there
were no associations between cord blood Pb and pubic hair or breast development at age 13. In a cohort
of Mexican children, cumulative blood Pb from 1-4 years old was associated with delayed breast
development (OR: 0.96 [95% CI: 0.92, 0.99]) and delayed pubic hair development (OR: 0.95 [95% CI:
0.92, 0.99]), but maternal patella Pb and maternal tibia Pb were not associated with either breast or pubic
hair development in girls between 9-18 years old (n = 283) (Liu et al.. 2019b). Additionally, the highest
tertile of maternal patella Pb and the second tertile of cumulative blood Pb from 1-4 years of age was
associated with delayed menarche. In a subset of the ELEMENT project cohort (n = 200), maternal blood
was measured during each trimester of pregnancy and daughters (mean age at follow-up assessment
13.8 ± 2.0 years) were asked about the occurrence of their first menstrual cycle (Jansen et al.. 2018). Only
second trimester maternal BLLs were associated with later age at menarche (HR: 0.59 [95% CI: 0.28,
0.90]).
8.4.2.2 Toxicological Studies on Puberty among Females
There were no recent animal toxicological studies on the effects of Pb on puberty in females. The
2013 Pb ISA reported that one research group Iavicoli et al. (Iavicoli et al.. 2006; Iavicoli et al.. 2004)
observed that mouse offspring exposed to Pb prior to birth and through puberty (BLLs 0.7-13 (ig/dL)
resulted in a dose-dependent delay in multiple markers of sexual maturation (e.g., vaginal opening, age at
first estrus). The latter study by this group utilized a multigenerational dosing paradigm in which they
observed delays of pubertal onset in the F2 generation similar to those seen in the Fi. The 2013 Pb ISA
also summarized a study in which Fisher 344 rats were dosed daily via gavage (12 mg/mL Pb) starting
30 days prior to breeding through weaning of the offspring (PND 23), resulting in delayed age at vaginal
opening in the offspring (BLLs of dams just prior to breeding averaged 39.8 (ig/dL) (Pine et al.. 2006). Of
particular interest is that the observed delay in vaginal opening was attenuated in offspring that received
IGF-1 injections starting on PND 28 until vaginal opening was observed, demonstrating that IGF-1 is a
critical element to Pb-induced pubertal onset delays. Reports of delayed puberty due to Pb exposure in the
2013 Pb ISA are consistent with studies in the 2006 Pb AQCD which observed delays in puberty in
female Fisher 344 and Sprague-Dawley rats exposed to Pb during gestation and/or lactation (Dearth et al..
2004; Dearth et al.. 2002; Ronis et al.. 1998c; Ronis et al.. 1996).
8.4.2.3 Integrated Summary of Effects on Puberty among Females
There were several markers of puberty among females that were assessed for associations with Pb
exposures in the recent epidemiologic studies. Multiple cross-sectional studies and a single cohort study
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reported an association between blood Pb and delayed menarche among females, with relevant BLLs
(Gomula et al.. 2022; Jansen et al.. 2018; De Craemer et al.. 2017; Slawinska et al.. 2012). While these
cross-sectional studies reported imprecise associations, the pattern of association is important to note.
There was also indication of slower breast development, but the studies assessing breast development
were limited (Nkomo et al.. 2018; De Craemer et al.. 2017). Additionally, there were a limited number of
studies that evaluated the development of pubic hair, but these results were inconsistent (Liu et al.. 2019b;
Nkomo et al.. 2018; De Craemer et al.. 2017). A single study among NHANES participants reported
increased serum total testosterone levels in female adolescents. The recent epidemiologic studies
assessing the associations between Pb exposure and puberty among females were limited by the timing of
the exposure to Pb and biomarker of exposure (blood, maternal bone, cord blood). However, these studies
consider a wide range of confounders, including height, weight, and BMI, and the associations reported
demonstrate that Pb exposure could result in physiological responses that contribute to changes in puberty
in females.
No recent PECOS-relevant toxicological studies investigated the effects of Pb on puberty.
However, studies from the 2013 Pb ISA and the 2006 Pb AQCD provide toxicological evidence that
indicates that Pb delays onset of puberty in female rodents. Several studies report delays in pubertal
markers such as vaginal opening and first estrus. Of note is that one study, Pine et al. (2006) reported that
the observed delay in vaginal opening in Pb-treated animals was attenuated when Pb treated animals were
supplemented with IGF-1 starting on PND 28. This strongly suggests that the mechanism through which
Pb induces delays pubertal onset in females is dependent on IGF-1 disruption.
8.4.3 Effects on Puberty among Males
The recent epidemiologic and toxicological studies examining the relationship between Pb
exposure and effects on puberty among males are summarized in the text below. Study details of the
recent epidemiologic studies are included in Table 8-12.
8.4.3.1 Epidemiologic Studies on Puberty among Males
The epidemiologic studies reviewed in the 2013 Pb ISA demonstrated an inverse relationship of
Pb on pubertal development among males at low blood Pb (mean and/or median BLLs of 3.0 to
9.5 (ig/dL). The studies were mostly cross-sectional, but the findings from these studies were supported
by those from a prospective longitudinal study (Williams et al.. 2010). Boys with higher (>5 (ig/dL) BLLs
at ages 8-9 years old had 24% to 31% reduction of pubertal onset based on testicular volume (TV),
genitalia staging, and pubic hair staging (Williams et al.. 2010). While temporality of effects is difficult to
establish due to the nature of the cross-sectional study design, the larger studies controlled for potential
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confounders, with a few studies considering the inclusion of dietary factors, but did not control for other
metal exposures that may impact the associations.
In an NHANES (2011-2012) analysis, concurrent BLLs and serum total testosterone levels were
measured in 6-19-year-old children and adolescents (Yao et al.. 2019). Testosterone is a principal sex
hormone needed for the normal physiologic processes during all life stages. In males, testosterone is
essential for the development and maintenance of secondary sexual traits, and can also influence bone
mass, muscle strength, mood, and intellectual capacity. When comparing the highest quartile of blood Pb
to the lowest, there was no association between serum testosterone levels in either male children (|3:
-13.09% [95% CI: -34.45%, 15.22%]) or male adolescents (|3: 6.32% [95% CI: -14.62%, 32.4%]).
Among three successive, cross-sectional Flemish Environment and Health Studies (FLEHS I, FLEHS II
and FLEHS III) of adolescents in Belgium between 2002-2015, blood Pb was negatively associated with
free estradiol (fE2; OR: 0.908 [95% CI: 0.839, 0.983]) and free testosterone (OR: 0.909 [95% CI: 0.828,
0.997]) in FLEHS II, but not in FLEHS I (De Craemer et al.. 2017). The associations between blood Pb
and estradiol (E2), testosterone, sex hormone binding globulin (SHBG), luteinizing hormone (LH), and
follicle stimulating hormone (FSH) were generally null (see Table 8-12). In addition to sex hormones, De
Craemer et al. (2017) also evaluated the associations between blood Pb and genital development and
pubic hair development among male adolescents. Across the three cross-sectional studies, there was
decreased odds of delayed onset of genital development (FLEHS I OR: 0.843 [95% CI: 0.717, 0.99];
FLEHS II OR: 0.697 [95% CI: 0.462, 0.998]; FLEHS III OR: 0.621 [95% CI: 0.388, 0.967]) and pubic
hair development (FLEHS I OR: 0.808 [95% CI: 0.686, 0.949]; FLEHS II OR: 0.849 [95% CI: 0.563,
1.365]; FLEHS III OR: 0.515 [95% CI: 0.327, 0.774]).
In the BT20+ birth cohort in South Africa, cord BLLs and blood levels at age 13 were evaluated
in association to puberty progression pubic hair development and genital development in 732 males
(Nkomo et al.. 2018). In males, elevated cord BLLs (>5 (ig/dL) was associated with slower pubic hair
development (RR: 0.28 [95% CI: 0.11, 0.74]). There were no associations between cord blood Pb and
genital development or BLLs at age 13 with pubic hair development or genital development. Similarly,
there were no associations between maternal patella Pb, maternal tibia Pb, and cumulative blood Pb from
1-4 years old and pubertal development (genitalia, pubic hair, and TV) in boys (Liu et al.. 2019b).
However, in a longitudinal cohort of boys from the Russian Children's Study, higher BLLs
(>5 (ig/dL) measured at age 8-9 years old (baseline) had pubertal onset 7.7-8.4 months later, on average,
than those with lower BLLs (<5 (ig/dL) (Williams et al.. 2019). Boys with higher BLLs at baseline had
later adjusted mean age at sexual maturity, with 4.2-5.1 months later attainment compared to boys with
lower BLLs. There was a shift in mean age for age at pubertal onset for stage 2 genitalia (G2) of
8.40 months (95% CI: 3.70, 13.10), 8.12 months (95% CI: 3.46, 12.78) for stage 2 pubic hair (P2), and
7.68 months (95% CI: 3.46, 11.90) for TV >3 mL. There was a shift in mean age for age at sexual
maturity for stage 5 genitalia (G5) of 4.20 months (95% CI: 0.56, 7.84) and 5.14 months (95% CI: 1.70,
8.58) for TV >20 mL, but a null association for stage 5 pubic hair (4.23 months [95% CI: -0.31, 8.77]).
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Furthermore, there was no shift in the mean age for duration of pubertal progression for genitalia (G2 to
G5), pubic hair (P2 to P5), or TV (>3 mL to >20 mL). In a mediation analysis, growth measurements at
age 11 were included to better understand what portion of the shift in mean age at sexual maturity was
attributable to the effect of BLL on growth. The association of peripubertal BLL with height Z-score
(HTZ) at age 11 accounted for 34%-53% of the total effect of BLLs on age at maturity, while BMIZ-
score at age 11 only accounted for 6%-23%. In another Russian Children's Study, Fleisch et al. (2013)
longitudinally measured serum insulin-like growth factor (IGF-1) to assess the association with childhood
BLLs. BLLs were measured at baseline only and IGF-1 levels were only measured during the follow-up
periods. Boys were enrolled between the ages of 8-9 years and were prospectively followed, with IGF-1
measurements obtained at two-year follow-up (ages 10-11 years) and at four-year follow-up (ages 12-
13 years). The overall mean IGF-1 concentration was 29.2 ng/mL lower (95% CI: -43.8, -14.5) for boys
with high BLLs at age 8-9 years (>5 (ig/dL [max is 31 |ig/dL|) versus those with lower baseline BLLs
(<5 (ig/dL) in adolescence among boys.
8.4.3.2 Toxicological Studies on Puberty among Males
There were no recent toxicological studies on the effects of Pb on puberty in males, as was also
the case at the time of the 2013 Pb ISA. The 2006 Pb AQCD reported that one study found that prenatal
exposure to Pb delayed sexual maturation in a dose-dependent manner in male rats (Ronis et al.. 1998c).
Specifically, Ronis et al. (1998c) reported that prostate weight (used in this study as a marker of sexual
maturation) was reduced in male rat offspring treated with Pb from GD 5 through sacrifice.
8.4.3.3 Integrated Summary of Effects on Puberty among Males
The epidemiologic studies reviewed in the 2013 Pb ISA demonstrated an inverse relationship of
Pb on pubertal development among males at low blood Pb (mean and/or median BLLs of 3.0 to
9.5 (ig/dL). Overall, the recent epidemiologic studies assessing the associations between Pb exposure and
different markers of puberty (hormone levels, pubic hair development, genital development, TV) among
males reported more inconsistent findings at low BLLs. Additionally, there were differences in the timing
of exposure to Pb and different biomarkers of Pb exposure (maternal blood, maternal bone, cord blood, or
concurrent blood). The recent epidemiologic studies were able to consider a wide range of confounders,
including height, weight, and BMI, and some studies were conducted among established longitudinal
cohorts. No recent toxicological studies reported on the effects of Pb on male puberty. Similarly, the 2013
Pb ISA reported no studies investigating Pb and male puberty. One study reported by the 2006 Pb AQCD
investigated the effects of Pb on puberty using prostate weight as a marker of sexual maturity in male rats
exposed to Pb starting on GD 5. Treatment with Pb resulted in reductions in prostate weight around the
time of puberty, possibly indicating that Pb delayed the onset of sexual maturation in Pb-treated animals
when compared to control.
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8.4.4
Other Developmental Effects
There were several recent studies that evaluated associations between Pb exposure and other
developmental effects in the epidemiologic and toxicological literature. Study details of the recent
epidemiologic studies are included in Table 8-13 and the recent toxicological studies are in Table 8-11.
8.4.4.1 Epidemiologic Studies on Other Developmental Effects
There were several recent studies with other outcomes related to developmental effects. In studies
of other related to developmental effects, there was a negative association with child blood Pb and
mitochondrial DNA copy number (Alcgria-Torrcs et al.. 2020); positive associations with maternal Pb
exposure and diurnal Cortisol rhythms in infants (Tamavo v Ortiz et al.. 2016); and lower salivary sialic
acid levels (a metric for oral anti-inflammatory potential which may increase the risk of dental caries)
(Hou et al.. 2020). but no associations between tooth Pb levels (second trimester, third trimester, or
postnatal) and alpha diversity metrics (bacterial or fungal), indicators of gut microbiota (Sitarik et al..
2020) or child blood Pb and telomere length (Alegria-Torres et al.. 2020).
8.4.4.2 Toxicological Studies on Other Developmental Effects
The 2013 Pb ISA and the 2006 Pb AQCD did not report any studies that investigated the effects
of Pb exposure on developmental milestones. However, some recent studies have investigated these
outcomes (Table 8-11). One study dosed Wistar dams via drinking water (0.2% Pb) either prior to
conception, during gestation only, during lactation only, or during both gestation and lactation and
reported that only exposure during both gestation and lactation elicited impacts on developmental
milestones (Rao Barkur and Bairv. 2016). Specifically, the age at eye opening was reduced. However,
although this exposure paradigm was the only one that produced effects on age at eye opening, it was also
the only paradigm that resulted in BLLs higher than 30 (ig/dL and reported a BLL of 31.59 (ig/dL in pups
on PND 22. Rao Barkur and Bairv (2016) also investigated other developmental milestones such as pinna
detachment and tooth eruption but reported no Pb-induced changes in either of these outcomes. An
additional study investigated the effects of Pb on similar outcomes including eye opening, eye slit
formation, fur development, tooth eruption, and pinna detachment, but reported no effects when Wistar
dams were dosed via drinking water (0.2% Pb) from GD 6 to 21 (BLLs 11.2 (ig/dL in pups on PND 21)
(Basha and Reddv. 2015).
8.4.4.3 Integrated Summary of Other Developmental Effects
The recent epidemiologic studies have the potential to provide initial support of potential
mechanistic pathways for diurnal Cortisol rhythms, lower salivary sialic acid levels, and DNA oxidative
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stress damage from Pb exposure among children during developmental periods. However, the small
number of studies limits the ability to judge coherence and consistency across the outcomes evaluated in
these studies, although the associations with diurnal Cortisol rhythms, lower salivary sialic acid levels, and
decrease in mitochondrial DNA copy number indicate that Pb exposure could result in physiological
responses that may contribute to adverse developmental effects. Recent toxicological studies on other
developmental effects of Pb largely pertain to the effects of Pb on developmental milestones of offspring.
Of the few toxicological studies available, no effects of Pb on developmental milestones were reported
with the only exception being a reduction at the age of eye opening, but this treatment group had BLLs
higher than 30 (ig/dL.
8.5 Effects on Female Reproductive Function
The 2013 Pb ISA concluded that the relationship observed with female reproductive outcomes,
such as fertility and hormone levels in some epidemiologic and toxicological studies was sufficient
evidence to conclude a suggestive causal relationship between Pb exposure and female reproductive
function. Epidemiologic studies provided information on different exposure periods and support the
conclusion that Pb possibly affects at least some aspects of female reproductive function. However,
toxicological studies were less supportive for suggesting a causal relationship between Pb exposure and
female reproductive function. This may primarily be due to a lack of variety in female reproductive
endpoints investigated by studies identified in the literature. The only outcomes reported by PECOS-
relevant toxicological studies include litter size, number of litters, and maternal body weight.
Subsequently, no evidence was available for outcomes such as cyclicity, female hormones, sex organ
histopathology (including ovarian follicular counts), or female fertility indicators (e.g., latency to
pregnancy, implantation counts, conception rate). Additionally, the available toxicological evidence was
inconclusive, and the only studies that reported effects on female reproductive outcomes also reported Pb-
induced reductions in brain weight, indicating the possibility that animals were experiencing overt
toxicity from Pb (Saleh et al.. 2019; Saleh et al.. 2018). The recent epidemiologic and toxicological
studies of Pb exposure and female reproductive function are detailed in the following sections.
8.5.1 Effects on Hormone Levels and Menstrual/Estrous Cycle
The recent epidemiologic and toxicological studies examining the relationship between Pb
exposure and hormone levels and menstrual/estrous cycle are summarized in the text below. Study details
of the recent epidemiologic studies are included in Table 8-14 and the recent toxicological studies are in
Table 8-15.
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8.5.1.1 Epidemiologic Studies on Hormone Levels and Menstrual/Estrous Cycle
The epidemiologic studies reviewed in the 2013 Pb ISA reported associations between
concurrent/closely timed BLLs and hormone levels in female adults. However, while there were changes
in hormone levels, there were inconsistencies in the hormones that were evaluated across the different
studies. A limitation of some the epidemiologic studies evaluated was the cross-sectional design, which
leaves uncertainty regarding Pb exposure magnitude, timing, duration, and frequency that contributed to
the observed associations. Additionally, the covariates included in statistical models as potential
confounders varied among studies, which could contribute to between study heterogeneity. Another
limitation of the epidemiologic studies is that not all of the studies investigated important confounders,
such as other metal exposures or smoking. The recent epidemiologic studies are divided into studies on
hormone levels and studies on menstrual/estrous cycle. The recent epidemiologic studies on hormone
levels in the following section are specific to hormones related to reproductive function and recent
epidemiologic studies on other hormones are described in Section 9.4.2 in the Other Health Effects
Appendix.
8.5.1.1.1 Epidemiologic Studies on Hormone Levels in Females
There were a few cross-sectional studies that evaluated the associations between Pb exposure and
different hormones in females (Lee et al.. 2019; Chen et al.. 2016; Krieg and Feng. 2011). These studies
used population-based surveys to evaluate associations between blood Pb and hormones and found
consistent positive associations with FSH in post-menopausal women. In an NHANES (1999-2002)
analysis, Krieg and Feng (2011) evaluated serum FSH and LH. Serum FSH slope increased per every
logurblood Pb increase in the post-menopausal women (|3: 26.38 [95% CI: 13.39, 39.38]), women who
had both ovaries removed (|3: 27.71 [95% CI: 1.64, 53.78]), and pre-menopausal women (|3: 11.97 [95%
CI: 3.27, 20.66]), but serum FSH was not associated with BLLs in pregnant women, women who were
menstruating, or women who were taking birth control pills. Serum LH slope increased per every logur
blood Pb increase in the post-menopausal women (|3: 11.63 [95% CI: 4.40, 18.86]) and women who had
both ovaries removed (|3: 20.59 [95% CI: 2.14, 39.04]), but serum LH was not associated with BLLs in
the pregnant women, women who were menstruating, women who were taking birth control pills, and
pre-menopausal women. In another cross-sectional, population-based survey in China, Chen et al. (2016)
examined the associations between blood Pb and total testosterone (tT), E2, and SHBG, in addition to LH
and FSH in postmenopausal women (age >55 years). When comparing the highest quartile of blood Pb
(>5.98 (ig/dL) to the lowest (<2.70 |ig/dL). there were positive associations with BLLs and SHBG (|3:
0.048, SE: 0.016, p < 0.01), FSH (|3: 0.046, SE: 0.016, p < 0.01), and LH (|3: 0.037, SE: 0.016, p < 0.05).
There were null associations between BLLs and tT or E2. Across the quartiles of blood Pb, there were
also positive trends observed with SHBG (p for trend: 0.002), FSH (p for trend: 0.001), and LH (p for
trend: 0.026), suggesting a potential linear exposure response between blood Pb and these hormones. In a
study of the Korea National Health and Nutrition Examination Survey (KNHANES) (2012-2014), blood
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Pb and serum FSH levels were assessed in postmenopausal women (aged 50 or older) (Lee et al.. 2019).
Serum FSH levels were positively associated with increasing blood log-Pb (|3: 2.929 [95% CI: 0.480,
5.377]).
8.5.1.1.2 Epidemiologic Studies on Menstrual/Estrous Cycle
There were no available epidemiologic studies in the 2013 Pb ISA that evaluated Pb exposure
with menopause. There are a limited number of studies that examined the relationship between Pb
exposure and menopause, and these recent studies reported consistent positive associations. A recent
cross-sectional study, NHANES (1999-2010) data was used to examine the associations between blood
Pb and menopause among women aged 45-55 years (age range where menopause is likely to occur)
(Mendola et al.. 2013). In the overall study sample (NHANES 1999-2010), with increasing quartiles of
blood Pb, there were increasing odds of menopause. Comparing the lowest BLLs (<1.0 (.ig/dL). the odds
for Q2 through Q4 were 1.7 (95% CI: 1.0, 2.8), 2.1 (95% CI: 1.2, 3.6), and 4.3 (2.6, 7.2), respectively.
When adjusting for bone measurements (either bone alkaline phosphatase or femoral neck bone density),
the associations were similar. In a subset (n = 434) of the Nurse's Health Study, the associations between
bone Pb levels and age at menopause were explored (Eum et al.. 2014). Compared with women in the
lowest tertile of tibia Pb (<6.5 |ig/g). those in the highest tertile (>13 |ig/g) were 1.21 years younger at
menopause on average (95% CI: -2.08, -0.35; p for trend: 0.006). Women in the highest tertile of tibia
Pb had an increased odds of 5.30 (95% CI: 1.42, 19.78; p for trend: 0.006) for early menopause
(menopause before age 45) compared with women in the lowest tertile. The associations with early
menopause were null across tertiles for patella Pb and BLLs.
8.5.1.2 Toxicological Studies on Hormone Levels and Menstrual/Estrous Cycle
There are no recent animal toxicological studies on the effects of Pb on the menstrual/estrous
cycle. The 2013 Pb ISA did not report any studies that investigated the effects of Pb on the
menstrual/estrous cycle, however some studies were summarized in the 2006 Pb AQCD. Specifically,
studies conducted in non-human primates found that Pb exposure increased menstrual cycle variability,
reduced days of menstrual flow, increased cycle length, and reduced progesterone (Franks et al.. 1989;
Laughlin et al.. 1987). However, another study with a lower BLL than the previous studies (<40 (ig/dL
versus 44-89 (ig/dL) did not report an effect on the menstrual cycle in a non-human primate species
(Foster et al.. 1992). Impacts of Pb on estrous cyclicity were examined in two rat studies that both utilized
multiple dosing paradigms to assess the varying impacts Pb exposure may have during different
developmental periods. Specifically, one study used the following exposure windows: gestation only,
lactation only, gestation and lactation, postnatal (from birth and continued past weaning), and continuous
(from the beginning of gestation continued past weaning) (Ronis et al.. 1998a). This study reported that
offspring in the postnatal and continuous exposure groups had fewer females who were regularly cycling.
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The other study was conducted by the same research group and utilized the following dosing windows:
post-pubertal (PND 60 to PND 74), pre-pubertal (PND 24 to PND 74), and in utero (GD 5 to PND 85)
(Ronis et al.. 1996). Rats in the pre-pubertal and in utero exposure paradigm groups experienced estrous
cyclicity disruption. While this study seems to indicate that pre-pubertal periods are more sensitive to
chemical insult, the previous study by Ronis et al. (1998a) suggests that normal cyclicity is recoverable
after cessation of exposure. However, it is worth noting that in both of these studies the BLLs were very
high, with a range of 63.2-264 (ig/dL for treatment groups that displayed treatment-related effects.
There are no recent animal toxicological studies on the effects of Pb on reproductive hormones.
The 2013 Pb ISA reported on a few studies that investigated the effects of Pb on reproductive hormones,
but none on cyclicity. Rodent studies reported that gestational and lactational exposure decreased
circulating levels of progesterone and E2 (Pillai et al.. 2010; Nampoothiri and Gupta. 2008). Dumitrescu
et al. (2008b) reported similar findings in adult female Wistar rats that were exposed to Pb for 6 months
via drinking water. Dumitrescu et al. (2008b) reported reductions in E2, progesterone, and FSH and
increases in LH and testosterone. The 2006 Pb AQCD reports findings from some toxicological studies
that show effects of Pb on hormones and cyclicity. Reductions in progesterone were observed in a study
wherein monkeys had BLLs of 25 to 30 (ig/dL, but no such reductions in progesterone were observed in
monkeys with even lower BLLs (10 to 15 (ig/dL) (Foster et al.. 1996).
8.5.1.3 Integrated Summary of Effects on Hormone Levels and Menstrual/Estrous
Cycle
The recent cross-sectional, population-based survey epidemiologic studies found consistent
positive associations between blood Pb and FSH in women who were post-menopausal. While these
studies are limited by their study design, the studies were conducted in well-established population-based
surveys. These studies considered a range of confounders, including controlling for BMI and smoking,
even co-exposure to other metals. The recent studies examining the relationship between menopause and
Pb exposure found consistent positive associations. The results from concurrent exposure of blood Pb
with menopause were supported by the results from a longitudinal cohort that examines bone Pb, a
cumulative biomarker of Pb exposure, and menopause, both the difference in age at menopause and risk
of early menopause. No recent PECOS-relevant toxicological studies reported on the effects of Pb on
hormone levels in females or menstrual or estrous cyclicity. However, previous toxicological evidence
suggests that Pb may disrupt reproductive hormones and menstrual and estrous cyclicity in females. Two
toxicological studies in rats reported disruptions in estrous cyclicity, and two toxicological studies based
in non-human primates reported alterations to different menstrual cycle aspects (e.g., length of cycle,
length of menstruation) and reproductive hormone levels. Additional rodent studies reported effects of Pb
on circulating reproductive hormone levels, including sex steroid hormones (progesterone, testosterone,
and E2) and gonadotropin hormones (LH and FSH).
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8.5.2
Effects on Female Fertility
Multiple epidemiologic and toxicological studies have examined the relationship between Pb and
female fertility. These studies are summarized in the text below. Study details of the recent epidemiologic
studies are included in Table 8-14 and the recent toxicological studies are in Table 8-15.
8.5.2.1 Epidemiologic Studies on Female Fertility
The epidemiologic studies reviewed in the 2013 Pb ISA examined a variety of fertility-related
endpoints. Although some studies demonstrated an association between higher Pb biomarker levels and
fertility/pregnancy, the results are inconsistent across studies. One limitation in most of these studies is
that the participants were women seeking help for fertility problems. The participants were not samples of
the general population and therefore cannot be generalized to all women of childbearing age. This may
also have introduced substantial selection bias into the study.
The recent epidemiologic studies also evaluated different outcomes to measure fertility. In an
NHANES (2013-2014 and 2015-2016) study, Lee et al. (2020) assessed whether BLLs were associated
with self-reported infertility by comparing BLLs of infertile women (n = 42) to pregnant women (n = 82).
There was increased risk of 2.60 (95% CI: 1.05, 6.41) per two-fold increase in BLLs of infertility. When
BLLs were categorized into tertiles, risk of infertility was more pronounced (OR: 5.40 [95% CI: 1.47,
19.78] interfile 2 (0.41-0.62 ^ig/dL) and OR: 5.62 [95% CI: 1.13, 27.90] interfile 3 (0.63-5.37 ^ig/dL),
respectively). In the LIFE Study, a cohort of couples were followed prospectively to assess persistent
environmental chemicals and human fecundity (Louis et al.. 2012). BLLs in both the female and male
partners were collected at baseline. Female BLLs were not associated with increased time to pregnancy in
the female exposure model (OR: 0.97 [95% CI: 0.85, 1.11]) or the couple exposure model (OR: 1.06
[95% CI: 0.91, 1.24]). However, there was decreased odds, or longer time to pregnancy, for male BLLs in
both the male exposure model (OR: 0.85 [95% CI: 0.73, 0.98]) and the couple exposure model (OR: 0.82
[95% CI: 0.68, 0.97]).
In a cross-sectional study among infertile women in Taiwan, Lai et al. (2017) examined the
associations between BLLs and diagnosis of endometriosis, which can cause infertility. Increasing tertiles
of BLLs was associated with higher odds of endometriosis (OR: 2.59 [95% CI: 1.11, 6.06] for T3
compared to Tl). In a cohort study, among couples undergoing the first in vitro fertilization (IVF) cycle,
maternal Pb levels were assessed with pregnancy outcomes (Li et al.. 2022). Pb levels in serum were
collected before oocyte retrieval. With increasing maternal serum Pb levels, there was a reduction in
successful implantation (OR: 0.85 [95% CI: 0.77, 0.94]) and clinical pregnancy (OR: 0.95 [95% CI: 0.91,
0.99]). Additionally, when maternal serum Pb was categorized into tertiles, there was a lower rate of
successful implantation (OR: 0.58 [95% CI: 0.40, 0.85]) in the highest Pb fertile, compared to the lowest
Pb fertile. Among tertiles of Pb serum levels, the associations were null for clinical pregnancy.
Furthermore, there was a negative association with maternal serum Pb and high-quality embryo rate (|3:
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-0.14 [95% CI: -0.32, -0.04]), but there were null associations with all other embryo quality indicators
(Table 8-14). In a cohort of 195 couples undergoing IVF, Pb was measured in serum and follicular fluid
from the female partner and semen from the male partner in association with six IVF outcomes (Zhou et
al.. 2021a). There were no associations between serum or follicular fluid Pb levels and any IVF
outcomes—normal fertilization, good embryo, blastocyst formation, high-quality blastocyst, pregnancy,
or live birth.
8.5.2.2 Toxicological Studies on Female Fertility
The 2013 Pb ISA reported some studies that investigated the effects of Pb on female fertility. A
handful of these studies reported that exposure to Pb reduced litter sizes in exposed female rats
(Dumitrescu et al.. 2008a; Teiion et al.. 2006) and mice (Iavicoli et al.. 2006). Contrasting this is a study
that found no changes in fertility rate or litter size in female rats treated prior to mating through pregnancy
(Nampoothiri and Gupta. 2008). Recent studies corroborate the findings of Nampoothiri and Gupta
(2008) and do not demonstrate any effects of Pb on female fertility in terms of litter size or number of
litters produced by dosed dams in mice (Schneider et al.. 2016; Corv-Slechta et al.. 2013) or rats (Rao
Barkur and Bairv. 2016; Barkur and Bairv. 2015; Weston et al.. 2014; Betharia and Maher. 2012). Among
these recent studies, a variety of dosing paradigms were utilized, including exposure during
preconception, lactation, gestation, and combinations thereof (BLLs ranged between 3.02-26.86 (ig/dL in
pups on PND 2-22). The contrast in effects on litter size between studies that do and do not report effects
of Pb on litter size is perplexing, and the inconsistencies of BLL measurements (e.g., some measured Pb
levels in offspring, some measured in dams, some studies did not report BLLs at all) between studies
further exacerbates the difficulty of reconciling these contrasts. However, some plausible explanations for
these differences exist and primarily involve differences in study design. Studies that reported reductions
in litter size due to Pb exposure tended to either use higher doses (Teiion et al.. 2006). longer dosing
durations (Iavicoli et al.. 2006). or dosed sires in addition to dosing dams (Dumitrescu et al.. 2008a) when
compared to studies that did not report reductions in litter sizes.
8.5.2.3 Integrated Summary of Effects on Female Fertility
Among the recent epidemiologic studies, there were inconsistent associations between Pb
exposure and female fertility. In studies among participants in the general population, there was an
increased risk of self-reported infertility and longer time to pregnancy. However, among studies with
women who were either seeking help at a fertility clinic or reported infertility, the associations were
inconsistent. Because the study participants included only a small sample of women who were either
seeking help at a fertility clinic or reported infertility, there may be selection bias and limits the
generalizability of the results as study participants have already been diagnosed and are seeking treatment
for fertility issues. Further, pregnancy outcomes, such as successful implantation, are more likely to be
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ascertained from women seeking treatment at fertility clinics. Additionally, the recent epidemiologic
studies were limited by the concurrently measured exposure and outcome, different biomarkers of
exposure (e.g., blood, serum, and follicular fluid), and a small number of participants. However, these
studies did include adjustment for potential confounders, including age, BMI, and partner exposure.
Previous toxicological evidence reported inconsistent effects of Pb on fertility in females. All recent
toxicological studies reported that female fertility was not affected by Pb exposure, even when a variety
of dosing paradigms were used.
8.5.3 Effects on Morphology and Histology of Female Sex Organs (Ovaries,
Uterus, Fallopian Tubes/Oviducts, Cervix, Vagina, and Mammary
Glands)
Recent epidemiologic and toxicological studies evaluating the association between Pb exposure
and morphology or histology of female sex organs (ovaries, uterus, fallopian tubes/oviducts, cervix,
vagina, and/or mammary glands) are limited. Study details for the single cross-sectional epidemiologic
study are included in Table 8-14 and the toxicological studies are included in Table 8-15.
8.5.3.1 Epidemiologic Studies of Morphology and Histology of Female Sex Organs
(Ovaries, Uterus, Fallopian Tubes/Oviducts, Cervix, Vagina, and Mammary
Glands)
In the 2013 Pb ISA, there were no epidemiologic studies available that evaluated Pb
concentrations and associations with morphology or histology of female sex organs (ovaries, uterus,
fallopian tubes/oviducts, cervix, vagina, and/or mammary glands). A recent cross-sectional study
examined the association between BLLs and rate of uterine fibroids and uterine fibroid volume (Ye et al..
2017). Among 288 (46 with fibroids and 242 without) pre-menopausal women included in the study,
there were null associations between blood Pb and the presence of uterine fibroids (OR: 1.39 [95% CI:
0.75. 2.56]) and volume of the largest fibroids (|3: 0.12 [95% CI: -2.26, 2.51]). When blood Pb was
categorized into quartiles, the association with volume of uterine fibroids remained null. While the
associations between blood Pb and rate of uterine fibroids and uterine fibroid volume were generally null,
the women with uterine fibroids had higher geometric mean BLLs than women without fibroids
(1.43 (ig/dL versus 1.35 (ig/dL, respectively).
8.5.3.2 Toxicological Studies of Morphology and Histology of Female Sex Organs
(Ovaries, Uterus, Fallopian Tubes/Oviducts, Cervix, Vagina, and Mammary
Glands)
There are no recent animal toxicological studies on the effects of Pb on morphology or histology
of female sex organs. The 2013 Pb ISA discussed a single study that reported Pb exposure increased
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membrane fluidity in granulosa cells in Charles Foster rats that were dosed via intraperitoneal injections
for 15 days (Nampoothiri and Gupta. 2006). The 2006 Pb AQCD also reported that exposure to Pb during
early pregnancy caused structural changes in the uterine epithelium in mice (Nilsson et al.. 1991; Wide
andNilsson. 1979).
8.5.3.3 Integrated Summary of Morphology and Histology of Female Sex Organs
(Ovaries, Uterus, Fallopian Tubes/Oviducts, Cervix, Vagina, and Mammary
Glands)
There was a single recent epidemiologic study evaluating associations between Pb exposure and
uterine fibroids. Although this was a small cross-sectional study, it was able to control for a large range of
confounders. The results from this single study are limited by the small sample size and concurrent
measurements of blood Pb and fibroids. Toxicological evidence regarding Pb exposure and female sex
organs is scarce. No recent PECOS-relevant toxicological studies were available. Previous studies
discussed in the 2013 Pb ISA and the 2006 Pb AQCD were scarce and reported few effects of Pb on
female sex organ morphology or histology.
8.6 Effects on Male Reproductive Function
The 2013 Pb ISA concluded that there was toxicological evidence with supporting epidemiologic
evidence to conclude that a causal relationship exists between Pb exposure and effects on male
reproductive function. The key evidence was provided by toxicological studies in rodents, non-human
primates, and rabbits showing detrimental effects on semen quality, sperm, and fecundity/fertility with
supporting evidence in epidemiologic studies of associations between Pb exposure and detrimental effects
on sperm. Recently published research has continued to support an association between Pb and
sperm/semen production, quality, and function. Studies of Pb and male reproductive function are
described in the sections below.
8.6.1 Effects on Sperm/Semen Production, Quality, and Function
Multiple epidemiologic and toxicological studies have examined the relationship between Pb and
sperm and semen production, quality, and function. These studies are summarized in the text below.
Study details of the recent epidemiologic studies are included in Table 8-16 and the recent toxicological
studies are in Table 8-17. The majority of the recent epidemiologic studies are cross-sectional with
concurrent measurements of Pb levels in biological samples and sperm-related outcomes. Recent
toxicological studies use a variety of dosing paradigms, and those that dose for longer periods of time
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(>30 days) or during a developmental window most often reported effects of Pb exposure on aspects of
sperm and semen quality.
8.6.1.1 Epidemiologic Studies on Sperm/Semen Production, Quality, and Function
The 2013 Pb ISA epidemiologic studies of Pb exposure and sperm and semen production, quality,
and function were cross-sectional, mostly in occupational cohorts, with concurrent measurements of Pb
levels in biological samples and sperm-related outcomes. The multiple epidemiologic studies in
occupational cohorts had mean BLLs over 40 (ig/dL for individuals occupationally exposed to Pb. The
occupational studies also had limited consideration for potential confounding factors, such as other
workplace exposures, which may impact the associations. The epidemiologic studies of men attending
fertility clinics may be subject to selection bias, and the results may not be generalizable. Additionally,
these studies reported imprecise estimates, did not control for other potential confounders such as other
metals, and had small sample sizes.
Several recent cross-sectional studies have explored the relationship between Pb exposure and
sperm and semen production, quality, and function. These studies were all conducted in males attending
fertility clinics and reported inconsistent associations for various metrics of sperm/semen production,
quality, and function. There were other cross-sectional studies that also examined associations with sperm
and semen production, quality, and function using different and Pb measured in semen, seminal fluid, and
seminal plasma, but these findings were more inconsistent.
Among the cross-sectional studies that evaluated associations with blood Pb, there was lower
normal sperm morphology with increasing BLLs (Shi et al.. 2021; Sukhn et al.. 2018; Li et al.. 2015).
Additionally, Li et al. (2015) reported increased odds of lower semen quality, sperm concentration,
numbers of sperm, total motility sperm, and progressive motility sperm with increasing BLLs, whereas
Sukhn et al. (2018) reported null associations with sperm volume, concentration, total count, progressive
motility, viability, and World Health Organization (WHO) morphology, and Shi et al. (2021) reported
null associations between blood Pb and semen parameters of semen volume, sperm concentration, total
sperm count, sperm motility, total motile sperm count, sperm vitality, DNA fragmentation index, and
percentage of acrosome reacted sperm (see Table 8-16). Furthermore, there were differences in BLLs
between men categorized as having low-quality semen and those classified as having normal or high-
quality semen (Sukhn et al.. 2018; Li et al.. 2015). Li et al. (2015) reported mean blood Pb for men in the
low-quality semen group was 3.43 (ig/dL and 2.38 (ig/dL for those in the high-quality semen group, and
Sukhn et al. (2018) reported the mean blood Pb for low-quality semen group of 5.198 (ig/dL and
3.575 (ig/dL for the normal-quality semen group.
Other cross-sectional studies examined the relationship between Pb measured in seminal fluid and
metrics of sperm/semen production, quality, and function, but the associations were inconsistent (Jia et
al.. 2022; Sukhn et al.. 2018; Pant et al.. 2014). Pant et al. (2014) measured Pb in semen and metrics of
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sperm/semen production, quality, and function, and reported associations of higher Pb levels in sperm
with decreased sperm motility (|3: -2.43% [95% CI: -4.87%, -0.001%]), decreased sperm concentration
(|3: 1.97 106/mL [95% CI: -3.16, -0.33]), increased tail length (|3: 3.79 [95% CI: 0.56, 7.02]), increased
percent DNA in tail (|3: 1.31 [95% CI: 0.17, 3.74]), and increased tail movement (|3: 1.20 [95% CI: 0.23,
2.16]). However, Sukhn et al. (2018) assessed sperm characteristic relationships with seminal fluid Pb and
reported increased odds of below-reference sperm viability and WHO morphology with higher seminal
fluid Pb, but null associations with volume, concentration, total count, and progressive motility. Jia et al.
(2022) reported no associations between seminal plasma Pb concentrations and semen parameters (semen
volume, sperm concentration, total sperm number, progressive motility, and normal morphological rate).
A recent cohort study examined the associations between peripubertal blood Pb, collected at
enrollment, and parameters of sperm and semen production, quality, and function for 223 participants in
the Russian Children's Study, with the semen sample collected 10 years after enrollment (Williams et al..
2022). There were null associations between peripubertal blood Pb and sperm parameters (sperm
concentration, total count, progressive motility, and total progressive motile sperm count, or probability
of having low semen quality based on sperm count/motility), whether blood Pb was modeled
continuously, categorized as tertiles, or categorized as low (<5 (ig/dL) blood Pb versus high (>5 (ig/dL)
blood Pb (see Table 8-16).
8.6.1.2 Toxicological Studies on Sperm/Semen Production, Quality, and Function
The 2013 Pb ISA summarized several toxicological studies that investigated the effects of Pb
exposure on sperm-related outcomes. Utilizing a variety of dosing paradigms and animal models,
previously published studies have demonstrated that Pb exposure reduced sperm counts, reduced numbers
of viable sperm, reduced motility, and increased morphological abnormalities (Pillai et al.. 2012; Anium
et al.. 2011; Allouche et al.. 2009; Oliveira et al.. 2009; Salawu et al.. 2009; Shan et al.. 2009; Tapisso et
al.. 2009; Massanvi et al.. 2007; Piao et al.. 2007; Wang et al.. 2006). Results from recently published
studies tend to suggest that Pb exposure impacts sperm and semen parameters (Table 8-17). All available
studies that reported outcomes on sperm and semen parameters were conducted in mice. Only one study
utilized a developmental exposure paradigm and dosed lactating CD-I mice from PND 0 to 21 which
resulted in reduced numbers of sperm at PND 70 in male offspring in the highest dose group (BLLs
19.1 (ig/dL) (Wang et al.. 2013a). Other studies that directly exposed male mice after weaning also
reported Pb-induced sperm alterations including increased incidence of abnormal morphology, reduced
density, and reduced viability (BLLs 9.4-11.92 (ig/dL) (Zhang et al.. 2021; Xie et al.. 2020; Godinez-
Solis et al.. 2019). However, some studies also reported no Pb-induced effects on sperm motility,
concentration, count, or viability (BLLs 9.4-11.8 (ig/dL) (Pavlova et al.. 2021; Godinez-Solis et al..
2019). It is worth noting that while Pavlova et al. (2021) reported no Pb-induced effects on sperm count,
Pb-treated animals had sperm counts 25% lower than those of control mice, but this effect failed to reach
statistical significance (p = 0.146). In terms of patterns in the reported data, studies that utilized long-term
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exposure (>30 days) or dosed during developmental periods tended to report effects of Pb on sperm or
semen parameters, and Pavlova et al. (2021) was the only study to use short-term exposure during
adulthood and also was the only study to report no effects at all on any sperm or semen parameters.
8.6.1.3 Integrated Summary of Effects on Sperm/Semen Production, Quality, and
Function
Among the recent epidemiologic studies that evaluated associations between Pb exposure
(measured in blood, semen, seminal fluid, or seminal plasma) and effects on sperm/semen production,
quality, and function, there were inconsistent findings, which was similar to the conclusion in the 2013 Pb
ISA. More consistent associations were observed for blood Pb with decreased sperm/semen production,
quality, and function than for Pb measured in semen, seminal fluid, or plasma; however, there are
limitations in the recent epidemiologic studies. All the cross-sectional studies were conducted in males
attending fertility clinics, which may have resulted in selection bias and limits the generalizability of the
results. Further, the small sample size from these cross-sectional studies also reduces the statistical power
to determine the precision of the associations. With concurrent measurement of Pb exposure with
outcomes related to sperm/semen production, quality, and function, temporality cannot be established.
Lastly, the use of different biomarkers (e.g., blood, semen, seminal fluid, or seminal plasma) to measure
Pb exposure and the different metrics of sperm/semen production, quality, and function limits the ability
to judge coherence and consistency across studies. Despite these limitations, it is important to note that a
wide variety of potential confounders were controlled for, including hormone levels, which could
potentially impact sperm/semen production, quality, and function. Previous and recent toxicological
studies generally reported that Pb alters some aspect of sperm or semen quality, such as sperm density,
motility, morphology, and viability, especially those studies that employed dosing during developmental
periods or for periods 30 days or longer. All recent toxicological evidence was produced from mouse
strains, but previous toxicological studies report similar effects in other species such as rats and rabbits.
8.6.2 Effects on Hormone Levels in Males
The epidemiologic and toxicological studies reviewed in the 2013 Pb ISA reported inconsistent
results regarding changes in hormone levels in men and associations with Pb exposure. The results from
the 2013 Pb ISA were similar to the findings from the 2006 Pb AQCD. Recent epidemiologic and
toxicological studies are reported below. Epidemiologic studies were mostly cross-sectional with blood
Pb measured concurrently with hormone levels. Study details for the epidemiologic studies, including
BLLs, study population characteristics, potential confounder, and select results, are in Table 8-16.
Previous toxicological evidence regarding the effect of Pb on hormones in males is somewhat
inconsistent, but most studies reported impacts of Pb on hormone levels. Recent toxicological studies are
extremely limited, but support previous toxicological studies reported in the 2013 Pb ISA that observed
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Pb-induced effects on hormones in males. Study details for the recent toxicological studies are in
Table 8-17.
8.6.2.1 Epidemiologic Studies on Hormone Levels in Males
In the 2013 Pb ISA, there were a few epidemiologic studies that evaluated hormone levels in
males in association with Pb exposure. The findings of these studies were inconsistent. The epidemiologic
studies were limited by their sample populations, often occupational cohorts or men recruited from
fertility clinics, which may not be representative of the general populations and limits the generalizability
of the results. More specifically, the occupational cohorts may have other metal exposures that were not
considered and may confound the associations, while studies conducted with subjects from fertility clinics
are subject to selection bias. While these studies included important confounders such as smoking, other
factors, such as exposure to other metals, were often absent. The cross-sectional study design of some of
the epidemiologic studies reviewed makes it difficult for temporality of effects to be established.
Additionally, most of the epidemiologic studies examined concurrent Pb exposure and hormone levels,
which may not reflect changes resulting from long-term exposures.
The recent epidemiologic studies on hormone levels detailed in this section are specific to
hormones related to reproductive function and recent epidemiologic studies on other hormones are
described in Section 9.4.2 in the Other Health Effects Appendix.
There are a few recent epidemiologic cross-sectional studies that evaluated the associations
between hormone levels in males and Pb exposure. There were consistent positive associations between
blood Pb and testosterone among these studies. One NHANES analysis combined three consecutive
cycles of NHANES (1999-2000, 2001-2002, and 2003-2004) to investigate the associations between
blood Pb and various sex hormones: testosterone, free testosterone, E2, fE2, androstenedione glucuronide,
and SHBG among men over 20 years old (Kresovich et al.. 2015). Comparing the highest quartile of
blood Pb exposure (>3.20 (ig/dL) to the lowest (<1.40 |ig/dL). testosterone was positively associated with
blood Pb (|3: 0.79, SE: 0.22) and there was an indication of exposure-response (p for trend: 0.0026). There
were null associations between blood Pb and all other sex hormones. In another NHANES (2011-2012)
study, blood Pb and serum testosterone were measured in men of reproductive age (18-55 years old)
(Lewis and Meeker. 2015). Of the 484 men included in the analysis, there was a 6.65% (95% CI: 2.09%,
11.41%) change in the serum testosterone concentration associated with a doubling (100% increase) in
blood Pb concentration. Chen et al. (2016) also reported positive associations between concurrent
testosterone and blood Pb concentrations. Utilizing data from a population-based survey, the Survey on
the Prevalence in East China for Metabolic Diseases and Risk Factors (SPECT)-China, 2,286 men were
included in the analysis to investigate the relationship between quartiles of BLLs and multiple
reproductive hormones - tT, SHBG, E2, LH, and FSH. When comparing the highest quartile of blood Pb
(>6.249 (ig/dL) to the lowest (<2.90 |ig/dL). there were positive associations with BLLs and tT (|3: 0.033,
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SE: 0.010, p < 0.01), SHBG (|3: 0.038, SE: 0.012, p < 0.01), FSH (|3: 0.030, SE: 0.015, p < 0.05), and LH
(|3: 0.028, SE: 0.013, p < 0.05), but null associations with E2 (|3: -0.003, SE: 0.017). Across the quartiles
of blood Pb, there were also positive trends observed with tT (p for trend: 0.012), SHBG (p for trend
<0.001), FSH (p fortrend <0.001), and LH (p for trend <0.001), suggesting apotential linear
concentration response.
While there were consistent positive associations between blood Pb and serum testosterone in the
cross-sectional studies, a single cohort study reported null associations. Among a subset of participants
(n = 453) in the Russian Children's Study, there were no associations between peripubertal BLLs and
hormones levels (testosterone, LH, or FSH) measured between 8-19 years of age, whether blood Pb was
modeled continuously or categorized (<5 (ig/dL versus >5 (ig/dL) (Williams et al.. 2022).
8.6.2.2 Toxicological Studies on Hormone Levels in Males
The 2013 Pb ISA discussed several studies that reported on the effects of Pb on hormone levels in
males. All studies were conducted in rats, and all directly dosed the tested animals save for one that
exposed gestating and lactating dams and measured hormones in offspring. Dosing durations varied from
21 days to 24 weeks, and most studies reported reductions in testosterone (Pillai et al.. 2012; Anium et al..
2011: Biswas and Ghosh. 2006: Rubio et al.. 2006). One study observed increased testosterone (Allouche
et al.. 2009) and another reported reductions in LH and FSH (Biswas and Ghosh. 2006). However, not all
studies observed effects on male hormones due to Pb exposure. One study observed no change in
testosterone (Salawu et al.. 2009) and another reported no change in FSH and LH levels despite reporting
increased testosterone levels (Allouche et al.. 2009). Only one recent PECOS-relevant toxicological study
was published that investigated the effects of Pb exposure on hormones in males (Table 8-17). This study
dosed nursing CD-I mice from PND 0 to 21 and reported reduced serum testosterone at weaning and
PND 70 and reduced testicular testosterone at weaning in offspring in the highest dose group (19.1 (ig/dL)
(Wang et al.. 2013a).
8.6.2.3 Integrated Summary of Effects on Hormone Levels in Males
The recent cross-sectional epidemiologic studies reported consistent associations between blood
Pb and testosterone; however, a single cohort study reported no associations. Of note, the recent cross-
sectional studies were in adult men, whereas the single cohort study was in male adolescents.
Additionally, the study by Chen et al. (2016) provides further support of a positive association with
SHBG, FSH, and LH and blood Pb. The positive trends among quartiles of blood Pb and testosterone,
SHBG, FSH, and LH provide insight on the possible concentration-response relationship. These studies
have robust sample sizes drawn from population-based surveys and controlled for a number of
confounders, including smoking, but the temporality of effects is difficult to establish due to the nature of
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cross-sectional study design. Recent toxicological evidence regarding the effects of Pb on hormones in
males is extremely limited, but in agreement with most studies summarized in the 2013 Pb ISA which
report effects of Pb on hormones in males.
8.6.3 Effects on Male Fertility
The recent epidemiologic and toxicological studies examining the relationship between Pb
exposure and male fertility are summarized in the text below. The epidemiologic studies on Pb exposure
and male fertility are limited. Previous epidemiologic studies were conducted among men seeking help at
fertility clinics. Study details of the recent epidemiologic studies are in Table 8-16. Previous and recent
toxicological evidence regarding the effect of Pb on male fertility is scarce, but generally reports reduced
fertility in males exposed to Pb. Study details of the recent toxicological studies are in Table 8-17.
8.6.3.1 Epidemiologic Studies on Male Fertility
The epidemiologic studies included in the 2013 Pb ISA that assessed associations between Pb
exposure and male fertility reported inconsistent findings. The few studies available for review were
conducted with cases that included men seeking help at fertility clinics, resulting in limited
generalizability of the studies because the study populations are not representative of the general
population. Additionally, by recruiting men who were seeking help at fertility clinics, there could be
selection bias, as their fertility status is already known and those seeking help at fertility clinics may be
different from men who have fertility issues who may be unaware of their condition and not seeking help
at a fertility clinic. Another study was conducted among occupationally exposed men, which may result in
differential exposures compared to the general population. Furthermore, another study did not control for
potential confounders.
There were a limited number of recent epidemiologic studies that examined associations between
Pb and male fertility. In the LIFE Study, a cohort of couples were followed prospectively to assess
persistent environmental chemicals and human fecundity (Louis et al.. 2012). BLLs in both female and
male partners were collected at baseline. While female blood Pb was not associated with increased time to
pregnancy, there was decreased odds, or increased time to pregnancy, for male BLLs in both the male
exposure model (OR: 0.85 [95% CI: 0.73, 0.99]) and the couple exposure model (OR: 0.82 [95% CI:
0.68, 0.97]). In a cohort of 195 couples undergoing IVF, Pb was measured in blood serum and follicular
fluid from the female partner and semen from the male partner in association with six IVF outcomes
(Zhou et al.. 2021a). There was a positive association between Pb in seminal plasma and the possibility of
obtaining agood embryo (RR: 1.86 [95% CI: 1.05, 3.11]), but the associations were null across all other
IVF outcomes (normal fertilization, blastocyst formation, high-quality blastocyst, pregnancy, or live
birth).
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8.6.3.2 Toxicological Studies on Male Fertility
Only a few studies on the effects of Pb on male fertility were summarized in the 2013 Pb ISA.
These studies reported that Pb-exposed males produced smaller litters and fewer implantations and
fetuses per dam (Anium et al.. 2011; Sainath et al.. 2011). Only a single recent study investigated fertility
outcomes in males exposed to Pb (Table 8-17). This study exposed ICR-CD-I mice from PND 91 to 136
via drinking water and reported that sperm from treated mice had reduced fertilization capacity, resulting
in fewer fertilized oocytes in vitro (9.4 (ig/dL) (Godinez-Solis et al.. 2019).
8.6.3.3 Integrated Summary of Male Fertility
Similar to the 2013 Pb ISA, there were only a few epidemiologic studies evaluating associations
between Pb exposure and male fertility and the findings were inconsistent. The results from these studies
are limited by the small sample size and the study population was recruited from a fertility clinic, which
may have resulted in selection bias and limits generalizability as the study population has already been
diagnosed and are seeking treatment for fertility issues. Further, male fertility related to pregnancy
outcomes, such as successful implantation and normal fertilization, are more likely to be ascertained from
couples seeking treatment at fertility clinics. Additionally, different biomarkers were used to assess Pb
exposure, as well as different metrics of male fertility across the studies. In terms of toxicological
evidence, previous and recent studies are few in number. However, all report a reduction of male fertility
in Pb-treated animals using outcomes such as litter size, implantations, and fertilized oocytes.
8.6.4 Effects on Morphology and Histology of Male Sex Organs
The toxicological studies in the 2013 Pb ISA supported historical findings that showed an
association between Pb exposure and changes in the sex organs as well as germ cells. There were no
epidemiologic studies available for review for the 2013 Pb ISA that examined the relationship between Pb
exposure and morphology or histology of male sex organs. The current epidemiologic and toxicological
studies examining the relationship between Pb exposure and effects on morphology and histology of male
sex organs are summarized in the text below with study details in Table 8-16 and Table 8-17,
respectively.
8.6.4.1 Epidemiologic Studies of Morphology and Histology of Male Sex Organs
In the 2013 Pb ISA, there were no epidemiologic studies available that evaluated Pb
concentrations and associations with morphology or histology of male sex organs. A recent cohort study
evaluated the associations between prenatal metal exposure and reproductive development in boys at 2-
3 years (Huang et al.. 2020). Serum concentrations of multiple metals, including Pb, were obtained from
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mothers in the Guangxi Birth Cohort Study throughout pregnancy, while reproductive development was
measured as TV and anogenital distance (AGD), categorized as anopenile distance (AGDap) and
anoscrotal distance (AGDas), in 2-3-year-old male children. When maternal serum Pb levels were
categorized by quartiles, infants in the highest quartile (serum Pb >1.23 |ig/L) had, on average, a
0.064 mL (95% CI: -0.124, -0.004) smaller TV, 0.060 cm (95% CI: -0.110, -0.011) shorter AGDap, and
0.115 cm (95% CI: -0.190, -0.039) shorter AGDas than infants in the lowest quartile (serum Pb <
0.54 |ig/L).
8.6.4.2 Toxicological Studies of Morphology and Histology of Male Sex Organs
This section is divided into the two main outcomes for the male sex organs: changes in weight of
male sex organs and changes in histology/morphology of male sex organs. The 2013 Pb ISA summarized
several studies that investigated the effects of Pb exposure on male sex organ weights. Several studies
reported decreases in weights of organs such as the testis, epididymides, vas deferens, seminal vesicles,
and prostate (Anium et al.. 2011; Pillai et al.. 2010; Dong et al.. 2009; Salawu et al.. 2009; Biswas and
Ghosh. 2006; Rubio et al.. 2006). The direction of effect was consistent, and any effects observed were
only decreases in organ weights. However, the 2013 Pb ISA noted that there were many other studies that
did not report effects on male reproductive organ weights even when using similar doses as those studies
that did observe effects, indicating that the impact of Pb on reproductive organ weights is somewhat
inconsistent. Recent studies have also investigated the effects of Pb exposure on male sex organ weight
(Table 8-17). Wang et al. (2013a) reported that dosing male CD-I mouse pups via their dams" drinking
water from PND 0 to 21 led to reduced absolute weight of testes in both treatment groups at weaning and
reduced relative testis weight in the highest treatment group at weaning (BLLs 19.1-21.2 (ig/dL on
PND 22 and 3.24-4.40 (ig/dL on PND 70). However, they observed no effect on the weight of the
prostate, seminal vesicle, or epididymides at weaning and no effects on relative weights of any
reproductive organ on PND 70. Similarly, another study reported that dosing Sprague-Dawley rats from
GD -10 to PND 183 had reduced absolute and relative testis weights (BLLs 18.6 (ig/dL) (Wang et al..
2013b). However, some studies reported that Pb did not alter the weights of testes or epididymides in ICR
mice (BLLs 6.02-21.66 (ig/dL) (Pavlova et al.. 2021; Satapathv and Panda. 2017).
The 2013 Pb ISA reported on studies that investigated the effects of Pb on the histopathology of
male sex organs in rodents exposed to Pb. One of the most common outcomes was alterations of
seminiferous tubule pathology, such as reduced length of some spermatogenic cycle stages within
seminiferous tubules, tubule damage, and tubule atrophy (El Shafai et al.. 2011; Shan et al.. 2009;
Massanvi et al.. 2007; Rubio et al.. 2006; Wang et al.. 2006). A few recent Pb studies have also reported
Pb-induced histopathological changes in male sex organs (Table 8-17). All recent studies were conducted
in mice, and exposure paradigms used between recent studies varied from developmental to exposure
only during adulthood. One study in CD-I mice that utilized developmental exposure (dosing dams from
lactational day 0 to 21) reported that Leydig cell numbers in the testes were reduced in the highest dose
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group at weaning and layers of spermatogenic cells within the seminiferous tubules were decreased in
both dose groups at weaning and PND 70 (BLLs at weaning 19.1-21.1 (ig/dL; BLLs at PND 70 3.24-
5.09 (ig/dL) (Wang et al.. 2013a). Some studies that dosed mice for 90 days following weaning reported
histopathological disruptions to the epididymal epithelial cells (BLLs 6.02-11.8 (ig/dL) (Xie et al.. 2020)
and that spermatogenic cells within seminiferous tubules were reduced in number (BLLs at 11.92 (ig/dL)
(Zhang et al.. 2021). Lastly, a study that dosed mice from PND 60 to 74 reported that the epithelium of
the seminiferous tubules was disorganized, the luminal region contained undifferentiated germ cells, and
some tubules had decreased diameter and germ cell number and displayed incomplete spermatogenesis
(BLLs 21.7 (ig/dL) (Pavlova et al.. 2021).
A few previous and recent studies concurrently investigated the effects of Pb on male sex organ
weight and histopathology, and effects within studies were coherent (Rubio et al.. 2006; Wang et al..
2013a; Pavlova et al.. 2021). Both Rubio et al.. 2006 and Wang et al.. 2013a reported reductions in male
sex organs as well as histopathological alterations, while Pavlova et al.. 2021 reported no effects of Pb on
sex organ weight or histopathology.
8.6.4.3 Integrated Summary of Morphology and Histology of Male Sex Organs
In the 2013 Pb ISA, there were no epidemiologic studies available that evaluated Pb
concentrations and associations with morphology or histology of male sex organs. A recent cohort study
reported decreased TV, shorter AGD, shorter anopenile distance, and shorter anoscrotal distance in 2-3-
year-old male children. While it is difficult to judge coherence and consistency from the findings of a
single study, this well-designed longitudinal cohort study does provide limited evidence of changes in
morphology and histology of male sex organs. Previous and recent toxicological studies are consistent in
reporting that Pb affects different aspects of sex organ histopathology. The most consistent effects appear
to be disruptions of histopathology of seminiferous tubules within the testes. However, there exists a data
gap regarding the effects of Pb on histopathology of other male sex organs such as the prostate,
epididymides, and seminal vesicles.
8.7 Biological Plausibility
This section describes the biological pathways that may underlie some reproductive and
developmental health effects from exposure to Pb. Figure 8-1 graphically depicts the proposed pathways
as a continuum of pathophysiological responses—connected by arrows—that may ultimately lead to the
observed delayed onset in both males and females and reduced sperm/semen production, quality, and
function. This discussion of how exposure to Pb may lead to these reproductive and/or developmental
events also provides biological plausibility for the epidemiologic results reported previously in this
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Appendix. In addition, most studies cited in this subsection are discussed in greater detail earlier in this
Appendix.
Disruption of
GnRH levels
and
hypothalamic-
pituitary-gonadal
Pb
Exposure
Delayed Pubertal
Onset in Males and
Females
Damage to
organs and
supportive
somatic cells
Reduced
Sperm/Semen
Production, Quality,
and Function
GnRH = gonadotropin-releasing hormone; Pb = lead.
Note: The boxes above represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and
the arrows indicate a proposed relationship between those effects. Solid arrows denote evidence of essentiality as provided, for
example, by an inhibitor of the pathway or a genetic knockout model used in an experimental study involving Pb exposure. Arrows
may connect individual boxes, groupings of boxes, and individual boxes within groupings of boxes. Progression of effects is
generally depicted from left to right and color-coded (gray, exposure; green, initial effect; blue, intermediate effect; orange, effect at
the population level or a key clinical effect). Here, population level effects generally reflect results of epidemiologic studies. When
there are gaps in the evidence, there are complementary gaps in the figure and the accompanying text below.
Figure 8-1 Potential biological pathways for reproductive and developmental
effects following exposure to Pb.
8.7.1 Pubertal Onset
When considering the available health evidence, plausible pathways connecting Pb exposure to
two health endpoints reported in epidemiologic and toxicological studies are proposed in Figure 8-1. The
first endpoint addressed in the figure above is delayed pubertal onset due to Pb exposure. Several
previous epidemiologic and toxicological studies that reported delays in pubertal onset in females
(Gollcnbcrg et al.. 2010; Naicker et al.. 2010; Dumitrescu et al.. 2008a; Iavicoli et al.. 2006; Denham et
al.. 2005; Selevan et al.. 2003; Wu et al.. 2003) and males (Williams et al.. 2010; Hauser et al.. 2008)
were summarized in the 2013 Pb ISA and several toxicological studies were summarized the 2006 Pb
AQCD (Pine et al.. 2006; Dearth et al.. 2004; Iavicoli et al.. 2004; Dearth et al.. 2002; Ronis et al.. 1998a.
1996). Some toxicological studies from the 2006 Pb AQCD also reported delays in pubertal onset in
males (Ronis et al.. 1998c; Sokol et al.. 1985). While no recent PECOS-relevant toxicological studies that
investigated the effects of Pb on pubertal onset were available, several recent epidemiologic studies
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reported associations between Pb exposure and delayed onset of puberty in males (Williams et al.. 2019;
Nkomo et al.. 2018; De Craemer et al.. 2017) and females (Gomula et al.. 2022; Jansen et al.. 2018;
Nkomo et al.. 2018; De Craemer et al.. 2017; Slawinska et al.. 2012). The proposed biologically plausible
pathway through which Pb induces delays in pubertal onset begins with the Pb-induced disruption of the
gonadotropin-releasing hormone (GnRH) levels which may occur through reduction of circulating IGF-1
levels. GnRH is a key hormone in the hypothalamic-pituitary-gonadal axis and hormonal signaling
pathways related to reproduction and pubertal onset. A recent epidemiologic study found negative
associations between BLLs at 8-9 years of age and IGF-1 in boys 2 and 4 years later (Fleisch et al.. 2013)
and toxicological studies have reported reduced IGF-1 levels and IGF-1R expression in the brains of
animals exposed to Pb (Li et al.. 2016; Li et al.. 2014; Dearth et al.. 2002; Ronis et al.. 1998b). IGF-1 is
known to act on GnRH neurons and affect GnRH secretion (Dees et al.. 2021; Daftarv and Gore. 2005).
which is responsible for the release of LH and FSH from the anterior pituitary, resulting in stimulation of
the gonads to begin producing sex steroid hormones and mature oocytes and spermatozoa. One
toxicological study conducted in female Fisher 344 rats found that Pb-induced delays of pubertal onset
could be reversed by supplementation with IGF-1 (Pine et al.. 2006). This study reported that
supplementation with IGF-1 also restored GnRH and LH levels in Pb-exposed rats, demonstrating that
IGF-1 disruption is a key component in delays in the onset of puberty mediated by Pb at BLLs at/above
35 (ig/dL.
Pb has also been shown in some in vitro studies to directly alter steroidogenic enzyme expression
(e.g., steroidogenic acute regulatory protein, 3|3-hydroxysteroid dehydrogenase, and aromatase) and levels
of sex steroid hormones important for proper sexual maturation, including progesterone, E2, and
testosterone (Huang and Liu. 2004; Srivastava et al.. 2004; Taupeau et al.. 2003; Huang et al.. 2002;
Thoreux-Manlav et al.. 1995). Additionally, although not all studies report relationships between Pb and
hormone levels, some epidemiologic studies have reported associations and some toxicological studies
have demonstrated effects of Pb exposure on steroidogenic enzymes and sex steroid hormones (Pollack et
al.. 2011; Tomoum et al.. 2010; Dumitrescu et al.. 2008b; Nampoothiri and Gupta. 2008; Telisman et al..
2007; Rubio et al.. 2006; Sokol et al.. 1985). Pb-induced disruptions of the hypothalamic-pituitary-
gonadal axis, steroidogenic enzymes, and their sex steroid products are plausible explanations for the
observed delays in pubertal onset reported in epidemiologic and toxicological studies.
8.7.2 Male Reproductive Function
The other health outcome proposed in Figure 8-1 is male reproductive function. Recent
epidemiologic studies have reported that Pb exposure is associated with reductions in a variety of semen
parameters, including sperm motility, sperm concentration, and normal sperm morphology (Shi et al..
2021; Sukhn et al.. 2018; Li et al.. 2015; Pant et al.. 2014). These findings are generally consistent with
the epidemiologic evidence presented in the 2013 Pb ISA (U.S. EPA. 2013). Further, toxicological studies
provide supporting evidence that Pb negatively impacts male reproductive function (see Section 8.6.1)
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(Anium et al.. 2011; Sainath et al.. 2011). Figure 8-1 shows a plausible biological pathway through which
Pb may act to reduce reproductive function in males.
The 2013 Pb ISA concluded that the evidence indicates a causal relationship between Pb
exposure and reduced quality of sperm, and that this relationship was likely mediated through the
generation of reactive oxygen species (ROS), leading to cellular damage (U.S. EPA. 2013). Specifically,
the 2013 Pb ISA summarized one study that reported Pb-induced increases in oxidative stress markers
and reductions in antioxidant enzyme levels in testicular plasma of rats (Salawu et al.. 2009). In addition,
several studies in the 2013 Pb ISA reported attenuation of Pb-induced reductions in sperm count, motility,
and viability when animals were co-administered substances with known antioxidant properties (Salawu
et al.. 2009; Shan et al.. 2009; Madhavi et al.. 2007; Rubio et al.. 2006; Wang et al.. 2006). Further
supporting the proposed pathway through which oxidative stress mediates Pb-induced effects are
additional studies that report that Pb exposure dysregulates antioxidant enzymes, leading to oxidative
stress and DNA damage in the affected tissues (Lopes et al.. 2016; Kagi and Vallee. 1960; Ommati et al..
In Press). Recent studies also support the proposed pathway and report an attenuation of Pb-induced
effects on aspects of male reproductive function (e.g., subfecundity, reduced sperm count) in animals
supplemented with antioxidants (Zhang et al.. 2021; Abdelhamid et al.. 2020; Alotaibi et al.. 2020; Naderi
et al.. 2020; Udefa et al.. 2020; Abdrabou et al.. 2019; Hassan et al.. 2019; Ommati et al.. 2019;
BaSalamah et al.. 2018; Hasanein et al.. 2018; Mabrouk. 2018; El Shafai et al.. 2011; Leiva et al.. 2011;
Sainath et al.. 2011; Ommati et al.. In Press). Although many studies report negative effects of Pb on
supporting somatic cells that have key functions in the spermatogenic cycle (e.g., Leydig cells, Sertoli
cells), Pb may also have negative effects directly on sperm cells. Direct contact of Pb with sperm cells has
been documented by multiple studies (Jia et al.. 2022; Sukhn et al.. 2018; Pant et al.. 2014). One recent
study reported that incubating sperm from healthy adult men for 4 hours with 30 |ig/m L or 8 hours with
either 15 or 30 |ig/mL Pb increased DNA fragmentation, possibly due to oxidative stress and Pb binding
to DNA phosphate residues, disrupting the process of chromatin condensation (Gomes et al.. 2015). In
another study, 4 hours of incubation of semen samples from healthy adult men with Pb reduced
intracellular levels of cyclic adenosine monophosphate (cAMP) (10, 50, and 100 |iM) and Ca2+ (2.5, 10,
50, and 100 (j,M), both of which are important in regulating sperm cell function (He et al.. 2016). In
support of this alternative mechanism of action is one non-PECOS relevant study (due to use of i.p.
injection route) that reported an attenuation of Pb-induced effects on reproduction in Pb-injected male
mice that were supplemented with CaCh (Golshan Iranpour and Kheiri. 2016). Disruption of intracellular
levels of key components such as cAMP and Ca2+is another way in which Pb can directly affect sperm
health and function outside of oxidative stress.
In summary, pathways are suggested by which Pb exposure can delay pubertal onset and reduce
sperm/semen production, quality, and function. Studies indicate that Pb exposure likely impacts the
hypothalamic-pituitary-gonadal axis in both males and females, leading to disruption of the onset of
puberty, a developmental period with increasing regard for its sensitivity to insult due to the vulnerability
of the various endocrinological events for which it is known. In addition, Pb exposure alters multiple
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aspects of male reproductive function. The production of adequate quantities of viable sperm is essential
for proper male fertility and reproduction. Pb exposure hampers this by negatively impacting both the
sperm cell and the supportive somatic cells that play key roles in the spermatogenic cycle through
increased oxidative stress and disruption of other important intracellular functions.
8.8 Summary and Causality Determination
The 2013 Pb ISA (U.S. EPA. 2013) made four causality determinations for Pb exposure and
(1) effects on pregnancy and birth outcomes; (2) effects on development; (3) effects on female
reproductive function; and (4) effects on male reproductive function. The 2013 Pb ISA concluded that the
evidence is suggestive of a causal relationship between Pb exposure and effects on birth outcomes; a
causal relationship between Pb exposure and effects on development, based on the findings of delayed
pubertal onset among males and females; suggestive of a causal relationship between Pb exposure and
effects on female reproductive function; and a causal relationship between Pb exposure and effects on
male reproductive function. The following sections detail the causality determinations based on the recent
epidemiologic and toxicological studies.
8.8.1 Summary of Effects on Pregnancy and Birth Outcomes
The 2013 Pb ISA concluded that based on the mix of inconsistent results of studies on various
birth outcomes and some associations observed in epidemiologic studies of preterm birth and low birth
weight/fetal growth, the evidence was suggestive of a causal relationship between Pb exposure and birth
outcomes. Some associations were observed between Pb and low birth weight in epidemiologic studies
that used postpartum maternal bone Pb or air Pb concentrations. Although associations were less
consistent for low birth weight with maternal blood Pb measured, during pregnancy or at delivery, or with
Pb measured in the umbilical cord and placenta (maternal blood Pb or umbilical cord and placenta Pb
were the biomarkers most commonly used in studies of low birth weight), some negative associations
between Pb biomarker levels and low birth weight or other measures of fetal growth were observed. The
effects of Pb exposure during gestation in animal toxicological studies included mixed findings, but most
studies reported reductions in birth weight of pups or birth weight of litters when dams were treated with
Pb. Thus, although evidence available was mixed, some associations observed in epidemiologic studies of
preterm birth and low birth weight or fetal growth provided suggestive evidence of a causal relationship
between Pb exposure and birth outcomes.
Compared to the evidence assessed in the 2013 Pb ISA, the evidence for associations between Pb
exposure and birth outcomes in the 2024 Pb ISA is notably stronger due to a recent quasi-experimental
study demonstrating decreased probability of preterm birth, decreased probability of low birth weight,
decreased probability of SGA, and increased birth weight (Bui et al.. 2022). Overall, among the recent
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epidemiologic studies, there was a pattern of elevated risk of preterm birth observed across several studies
from multiple geographic locations. Additionally, the recent epidemiologic studies of preterm birth
included populations for which mean/median maternal blood Pb values were below 10 (ig/dL and
controlled for wide range of confounders, including GA, other metals, and maternal health factors
(e.g., smoking, parity, BMI). There remain uncertainties regarding the critical window for the timing of
the exposure (e.g., during pregnancy, at delivery), biomarkers examined for Pb (e.g., maternal blood, cord
blood, maternal red blood cells, maternal serum, placental tissue), and evaluation of co-pollutants among
the epidemiologic literature with limited supportive evidence in the toxicological literature. While there
were no epidemiologic or toxicological studies examining Pb exposure and maternal health outcomes in
the 2013 Pb ISA, recent epidemiologic and toxicological studies reported inconsistent results regarding
maternal health outcomes and different maternal health outcomes were evaluated between the
epidemiologic and toxicological studies. Among the epidemiologic studies, there were consistent null
associations between maternal blood (blood, serum, and erythrocytes) Pb levels and GDM in studies that
reported mean/median blood Pb below 10 (ig/dL. Although some recent epidemiologic studies
investigated various pregnancy-related endpoints, the small number of studies limits the ability to judge
coherence and consistency across these studies. Among the few toxicological studies that investigated
maternal health, the only outcome reported was maternal weight gain during pregnancy. Most studies
reported no effects of Pb on maternal weight gain during pregnancy, and the few that reported reductions
in maternal weight gain also reported reductions in dam brain weight, a marker often indicative of overt
toxicity. This suggests that the observed reduction in maternal weight gain during pregnancy reported in
these studies may not be directly due to Pb exposure and may have been influenced by overt toxicity
experienced by the dams.
The recent epidemiologic and toxicological studies of birth outcomes reported inconsistent
findings overall. Among the recent epidemiologic studies of prenatal growth and Pb exposure, the
findings were inconsistent, and no effects on birth weight were reported in the recent toxicological
studies. The inconsistencies in the recent epidemiologic studies of prenatal growth and Pb exposure may
be due to differences in study design, the timing of the exposure, differences in biomarkers of exposure,
and the wide variation in prenatal growth outcomes assessed (birth weight, birth length, HC, GA). A few
studies were further limited by small sample size, which may cause imprecision in the measures of
association. Recent toxicological studies did not report any effects of Pb exposure on birth weight. Of
note is a previous study discussed in the 2013 Pb ISA that reported reduced litter weights at birth were
driven by reduced weights in female pups. No recent studies performed separate analyses of birth weight
for male and female pups, or they did not assess female pup weights at all. This suggests that the observed
lack of effects in recent literature could be due to a lack of sensitivity.
The recent epidemiologic studies of Pb exposure and birth defects, specifically NTDs, CHDs,
OFC defects and abdominal congenital malformations, reported inconsistent associations. While the
associations were generally null for Pb exposure (measured in placental tissue, umbilical tissue, maternal
blood serum, and umbilical cord serum) and NTDs, CHDs, and abdominal congenital malformations,
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there were positive associations with OFC defects when Pb was measured in placental tissue or maternal
blood. The small number of studies limits the ability to judge consistency and coherence across studies of
different birth defects (e.g., NTDs, CHDs, OFC defects, and abdominal congenital malformations), timing
of Pb exposure (e.g., second trimester, third trimester, and at delivery), differences in biomarkers
(e.g., placental tissue, umbilical tissue, maternal blood serum, and umbilical cord serum, maternal blood),
and confounders considered in the analyses. Additionally, the relatively small sample sizes in some
studies reduce the statistical power to determine the precision of the associations. Recent toxicological
studies report no effects of Pb on birth defects in offspring. This contrasts with some previously reviewed
studies that reported defects in offspring of Pb-exposed dams. However, dams in these previous studies
also experienced overt toxicity due to the high Pb doses used, which did not occur in recent toxicological
studies, suggesting that maternal toxicity may have been involved with the birth defects observed in
previous studies.
There were only a few recent epidemiologic studies that evaluated Pb exposure and spontaneous
abortion and pregnancy loss. Studies that examine spontaneous abortion are difficult to conduct as many
spontaneous abortions or pregnancy losses occur during the first trimester. Women may miscarry before
being enrolled in a study and/or women may not have known they were pregnant when they miscarried,
thus limiting the ability of a study to detect subtle effects (e.g., if higher Pb exposures lead to increased
risk of early spontaneous abortions). In the recent epidemiologic studies, some of the studies assessing
spontaneous abortion and/or pregnancy loss were among women who were undergoing treatment at
fertility clinics. Detection of spontaneous abortion and/or pregnancy loss is more likely to be ascertained
in such clinics, but this study design approach may result in selection bias and limited generalizability of
the results because the study populations are not representative of the general population as they have
already been diagnosed and are seeking treatment for infertility. In the recent toxicological studies, there
were no reported effects of Pb exposure on pre- or postnatal offspring mortality. Although not always
consistently so, BLLs were generally lower in recent toxicological literature when compared to previous
literature, possibly explaining the observed contrast in results.
There were no epidemiologic studies available that evaluated Pb concentrations and associations
with placental function in the 2013 Pb ISA. There were a limited number of recent epidemiologic studies
in this area. These cross-sectional studies provide insight into associations between concurrent Pb
exposure and placental function, but are limited by their cross-sectional design, making it difficult to
establish the temporality of the effects or the critical window of exposure to Pb that might result in
changes in the placenta during pregnancy. Further, there were only a small number of cases, which may
result in imprecise associations. While previous toxicological evidence included decreased placental
weight and histological alterations, these findings were limited to a single study. Recent toxicological
studies reported that dams dosed with Pb had reduced placental weight, but some of these studies also
reported reduced brain weight in dams, suggesting that overt toxicity may have occurred and could be
related to the observed reductions in placental weight.
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There were also a number of recent epidemiologic studies that evaluated other outcomes related
to maternal health during pregnancy such as biomarkers of fetal immune system, fetal marker for
metabolic function, and rTL, but the small number limits the ability to judge the coherence and
consistency across these studies. The only additional pregnancy outcome investigated in recent
toxicological literature was sex ratio of offspring born to Pb-treated dams. Although most toxicological
studies reported no effects of Pb on sex ratio, a single study reported that Pb produced female-skewed
litters when compared to control. It is worth noting, however, that the non-Pb-exposed groups were male-
skewed and had male:female offspring ratios of 1.4-1.5, whereas Pb-treated groups had male:female
ratios of 1 to 1.
In summary, the collective evidence is sufficient to conclude that there is likely to be a causal
relationship between Pb exposure and effects on pregnancy and birth outcomes. This determination
is largely driven by a recent quasi-experimental study that reported Pb-related changes in birth weight and
probability of low birth weight, preterm birth, and small for gestational age, in addition to other studies
demonstrating effects between Pb exposure and preterm birth (Table 8-1). Additionally, there were a few
high-quality epidemiologic studies that reported associations with relevant BLLs and prenatal growth,
birth defects, spontaneous abortion and pregnancy loss, and placental function, but the findings overall
were inconsistent. There is uncertainty related to exposure patterns resulting in likely higher past Pb
exposures, especially among maternal Pb levels. Additional uncertainties are related to biomarkers of
exposure (maternal blood, maternal serum, maternal bone, maternal erythrocytes, cord blood, cord blood
serum, placental tissue), the critical window of exposure, and co-pollutants confounding. Of note, the
cohorts in the recent epidemiologic literature would generally be expected to have had appreciable past
exposures to Pb; however, the extent to which adult BLLs in these cohorts reflect the higher exposure
histories is unknown as is the extent to which these past Pb exposures (magnitude, duration, frequency)
may or may not elicit effects on pregnancy and birth outcomes. The recent evidence from the
toxicological studies mostly reported no effects of Pb across pregnancy and birth outcomes. This may be
due to the exclusion of toxicological studies with BLLs greater than 30 (ig/dL, indicating the possibility
that most pregnancy and birth outcomes are only affected in laboratory animals at levels higher than most
environmentally relevant Pb exposure levels.
8.8.2 Summary of Effects on Development
The 2013 Pb ISA concluded that the collective body of evidence integrated across epidemiologic
and toxicological studies, based on the findings of delayed pubertal onset among males and females, was
sufficient to conclude that there is a causal relationship between Pb exposure and developmental effects.
Multiple epidemiologic studies of Pb and puberty in the 2013 Pb ISA showed associations between
concurrent BLLs and delayed pubertal onset for girls and boys. In cross-sectional epidemiologic studies
of girls (ages 6-18 years) with mean and/or median concurrent BLLs from 1.2 to 9.5 (ig/dL, consistent
associations with delayed pubertal onset (measured by age at menarche, pubic hair development, and
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breast development) were observed. In boys (ages 8-15 years), fewer epidemiologic studies were
conducted but associations between BLLs and delayed puberty were observed, including associations
among boys in a longitudinal study. These associations were consistently observed in populations with
mean or median BLLs of 3.0 to 9.5 (ig/dL. Potential confounders considered in the epidemiologic studies
of both boys and girls that performed regression analyses varied. Most studies controlled for age and
BMI. Other variables, such as measures of diet, socioeconomic status (SES), and race/ethnicity, were
included in some of the studies. Adjustment for nutritional factors was done less often and this could be
an important confounder. A study using a cohort of girls from NHANES controlled for various dietary
factors as well as other potential confounders and reported an association between increased concurrent
BLLs and delayed pubertal onset (Selevan et al.. 2003). A limitation across most of the epidemiologic
studies of BLLs and delayed puberty was the cross-sectional design, which does not allow for an
understanding of temporality. There was uncertainty with regard to the exposure frequency, timing,
duration, and level that contributed to the associations observed in these studies. Additionally, the
toxicological studies reviewed in the 2013 Pb ISA indicated that delayed pubertal onset may be one of the
more sensitive developmental effects of Pb exposure with effects observed after gestational exposures
leading to BLLs in the female pup of 1.3-13 (ig/dL (Iavicoli et al.. 2006; Iavicoli et al.. 2004). An
additional study reviewed in the 2013 Pb ISA reported increases in age at vaginal opening in Wistar rats
that were dosed prior to conception and in utero, but BLLs were not reported (Dumitrcscu et al.. 2008a).
These results are supported by studies reviewed in the 2006 Pb AQCD that reported delays in pubertal
onset in female rats and mice as measured by age at vaginal opening and age at first estrus (Pine et al..
2006; Dearth et al.. 2004; Dearth et al.. 2002; Ronis et al.. 1998a; Ronis et al.. 1998c; Ronis et al.. 1996).
BLL varied greatly between studies with some reporting effects occurring in dose groups with levels
below 30 (ig/dL (Dearth et al.. 2004; Dearth et al.. 2002). while others only report effects in groups with
BLLs higher than 30 (ig/dL (Ronis et al.. 1998a; Ronis et al.. 1998c; Ronis et al.. 1996). A key study
reviewed in the 2006 Pb AQCD, Pine et al. (2006) reported increased age at vaginal opening in Fisher
344 rats that was attenuated by supplementation of IGF-1. However, Pine et al. (2006) only reported
BLLs of dams (39.8 (.ig/dL). making it difficult to determine what BLLs in the offspring may have been
achieved to elicit such effects on puberty. Toxicological studies have also reported delayed male sexual
maturity as measured by sex organ weight, among other outcomes, seeing significant decrements at BLLs
of 20-34 (ig/dL (Ronis et al.. 1998c; Sokol et al.. 1985). Thus, the 2013 Pb ISA concluded that the data
from the toxicological literature and from epidemiologic studies demonstrated puberty onset in both
males and females was delayed with Pb exposure.
In the 2013 Pb ISA, findings from epidemiologic studies of the effect of Pb on postnatal growth
were inconsistent. Findings from the toxicological literature of the effect of Pb exposure on postnatal
growth summarized in the 2013 Pb ISA and the 2006 Pb AQCD were fairly consistent, and most studies
showed decreases in body weight of Pb-exposed offspring at postnatal time points, while one study
reported an increase in body weight at 1 year of age in male offspring only.
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The 2013 Pb ISA summarized some toxicological evidence that demonstrated the effect of Pb on
other developmental outcomes, including impairment of retinal development, effects on the lens of the
eye, and alterations in the developing hematopoietic, hepatic systems and teeth. No studies that
investigated more classic toxicological developmental milestones (e.g., eye slit formation, eye opening,
pinna detachment) were reported in the 2013 Pb ISA.
In the recent epidemiologic and toxicological literature, the relationships between Pb exposure
and puberty onset in both females and males, as well as postnatal growth, were reviewed. While there
were no recent PECOS-relevant toxicological studies in puberty in either females or males, the recent
epidemiologic studies reported consistent patterns of association between blood Pb exposure and delayed
age of menarche (Gomula et al.. 2022; Jansen et al.. 2018; De Craemer et al.. 2017; Slawinska et al..
2012) and some indication of slower breast development (Nkomo et al.. 2018; De Craemer et al.. 2017) in
females, which is similar to the findings from the epidemiologic studies reviewed in the 2013 Pb ISA.
However, the associations between Pb exposure and male pubertal onset were inconsistent among the
cross-sectional studies. The differences in markers of puberty in males (hormone levels, pubic hair
development, genital development, TV) may explain the inconsistencies in findings across recent studies.
While the studies assessing Pb exposure and female and male puberty were limited by differences in the
timing of exposure to Pb or Pb biomarker (blood, maternal bone, cord blood), these studies consider a
wide range of confounders, including height, weight, and BMI.
The recent toxicological and epidemiologic studies that evaluated the relationship between Pb
exposure and postnatal growth were inconsistent. The majority of recent toxicological studies did not
report changes in postnatal growth due to Pb exposure (Zhao et al.. 2021; Xie et al.. 2020; Rao Barkur
and Bairv. 2016; Basha and Reddv. 2015; Basgen and Sobin. 2014). However, some recent toxicological
studies reported decreases (Duan et al.. 2017; de Figueiredo et al.. 2014; Graham et al.. 2011) and
increases (Betharia and Maher. 2012) in body weight of offspring due to Pb exposure. Among
epidemiologic studies that evaluated the associations between blood Pb and postnatal growth in children
(older than 4 years) there were more consistent patterns of associations of decreased height and weight
(Signes-Pastor et al.. 2021; Kuang et al.. 2020; Zhou et al.. 2020; Deierlein et al.. 2019; Kerr et al.. 2019;
Choi et al.. 2017). Overall, there were negative associations between Pb exposure and specific postnatal
growth outcomes among the cross-sectional studies. However, among cohort studies, there were some
inconsistencies in the associations of Pb exposure and different postnatal growth outcomes. These
inconsistencies in the cohort studies may be due to differences in the timing of when Pb exposure was
measured, the biomarker of Pb exposure (maternal blood, maternal bone, cord blood, infant blood,
childhood blood), and the timing of the outcome. The current inconsistent findings of exposure to Pb and
postnatal growth are similar to those reported in the 2013 Pb ISA.
There was a small body of epidemiologic studies across various other developmental effects;
however, the small number of studies limits the ability to judge coherence and consistency across these
studies, although the associations reported demonstrate that Pb exposure could result in physiological
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responses that contribute to adverse developmental effects, including changes to diurnal Cortisol rhythms,
lower salivary sialic acid levels, and oxidative stress damage to DNA from Pb exposure among children
during developmental periods. Recent studies that investigate other developmental outcomes such as
developmental milestones are scarce. Some toxicological studies investigated developmental milestones
in rodents (e.g., pinna detachment, eye slit formation, eye opening, tooth eruption, and fur development),
but no effects of Pb exposure were reported on any of these milestones in groups with PECOS-relevant
BLLs.
In summary, the collective evidence is sufficient to conclude a causal relationship exists
between Pb exposure and effects on development. The key evidence is outlined in Table 8-1. While
there were no recent PECOS-relevant toxicological studies that investigated the impacts of Pb on puberty
in either females or males, previous toxicological evidence demonstrated that Pb exposure consistently
delayed pubertal onset in female rodents. Of note is one key previous toxicological study (Pine et al.,
2006) which demonstrated that delayed pubertal onset in female rats developmentally exposed to Pb
could be completely attenuated by supplementation with IGF-1. Further, the recent epidemiologic studies
reported consistent patterns of associations between blood Pb exposure and delayed age of menarche and
some indication of slower breast development in females, which is similar to the findings from the
epidemiologic studies reviewed in the 2013 Pb ISA. The few recent cohort studies of male pubertal onset
found consistent associations between Pb exposure and delayed onset. Recent cross-sectional studies
reported inconsistent results, possibly due to differences in the markers of puberty examined (hormone
levels, pubic hair development, genital development, TV). Though the effects of Pb exposure on postnatal
growth were inconsistent overall, there was some evidence from toxicological studies indicating reduced
body weight of offspring and from epidemiologic studies reporting associations between blood Pb and
decreased height and weight in children. The cohorts in the recent epidemiologic literature would
generally be expected to have had appreciable past exposures to Pb; however, the extent to which adult
BLLs in these cohorts reflect the higher exposure histories is unknown, as is the extent to which these
past Pb exposures (magnitude, duration, frequency) may or may not elicit developmental effects such as
decreased postnatal growth or disrupted puberty. Toxicological evidence supports biologically plausible
pathways of how Pb exposure exerts its effects on pubertal onset (Li et al., 2016; Li et al„ 2014; Pine et
al., 2006; Dearth et al., 2002; Ronis et al., 1998b), including studies suggesting that Pb may impact
pubertal onset via dysregulation of IGF-1 resulting in a cascade of effects that alters levels of hormones
important during the pubertal period.
8.8.3 Summary of Effects on Female Reproductive Function
The 2013 Pb ISA concluded that the available evidence was suggestive of a causal relationship
between Pb exposure and female reproductive function. Epidemiologic and toxicological studies of
reproductive function among females investigated whether Pb biomarker levels were associated with
hormone levels, fertility, menstrual/estrous cycle changes, and altered morphology or histology of female
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reproductive organs. Two previous toxicological studies conducted in non-human primates reported
disrupted menstrual cyclicity and reduced progesterone, although another non-human primate study with
lower BLLs than the other studies (<40 (ig/dL versus 44-89 (ig/dL) reported no effects on menstrual
cyclicity. Some previous toxicological studies in rodents also demonstrated impacts of Pb exposure on
estrous cyclicity, but of note is the high BLLs in treated animals (63.2-264 (ig/dL). Some of the
epidemiologic studies reviewed in the 2013 Pb ISA reported associations with concurrent BLLs and
altered hormone levels in adults, but results varied among studies, possibly due to the different hormones
examined and the different timing in menstrual and lifecycles. There was some evidence of a potential
relationship between Pb exposure and female fertility, but findings were mixed. The majority of the
epidemiologic studies were cross-sectional and adjustment for potential confounders varied from study to
study, with some potentially important confounders, such as BMI, not included in all studies. Further,
most of the epidemiologic studies on female reproductive function reviewed in the 2013 Pb ISA had
small sample sizes and were generally conducted in women attending infertility clinics. Previous
toxicological studies reported inconsistent effects of Pb on female fertility outcomes in rodents (reduced
litter size, reduced number of litters produced), while all recent toxicological studies reported no effects
of Pb exposure on female fertility outcomes. Studies that reported impacts of Pb on female fertility
outcomes tended to use higher doses, longer dosing durations, and/or concurrently exposed sires in
addition to dams, which may explain the observed contrast between studies. Although epidemiologic and
toxicological studies provide information on different exposure periods, both types of studies, including
some high-quality epidemiologic and toxicological studies, supported the conclusion that Pb may affect
some aspects of female reproductive function.
There were no recent PECOS-relevant toxicological studies of the effects of Pb exposure on
hormone levels in females or menstrual/estrous cyclicity; however, there were several recent
epidemiologic studies. The recent epidemiologic studies examining the relationship between Pb exposure
and hormone levels reported consistent positive associations between blood Pb and FSH and LH in
women who were post-menopausal. While these studies were limited by their cross-sectional study
design, the studies were conducted in well-established population-based surveys. These studies
considered a range of confounders, including controlling for BMI, smoking, and co-exposure with Cd, but
not all studies adjusted for some potential important confounders such as age at menarche, pregnancy
history, oral contraceptive use, and female hormone use, such as IVF or hormone therapy. Additionally,
the recent studies examining the relationship between menopause and Pb exposure found consistent
positive associations of early risk of menopause. The results from a study of concurrent exposure of blood
Pb with menopause were supported by the results from a longitudinal cohort that reported that bone Pb, a
cumulative biomarker of Pb exposure, was associated with difference in age at menopause and risk of
early menopause.
Among the recent epidemiologic studies, there were inconsistent associations between Pb
exposure and female fertility. In studies among participants in the general population, there was an
increased risk of self-reported infertility and longer time to pregnancy (Lee et al.. 2020; Louis et al..
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2012). However, among studies with women who were either seeking help at a fertility clinic or reported
infertility the associations were inconsistent. Because the study participants included only a small sample
of women who were either seeking help at a fertility clinic or self-reported infertility, selection bias may
exist and limits the generalizability of the results. Additionally, these studies were limited by the
concurrently measured exposure and outcome, different biomarkers of exposure (blood, serum, and
follicular fluid), and a small number of participants. These studies did include adjustment for potential
confounders, including age, BMI, and partner exposure. The recent toxicological studies in female
fertility did not observe alterations in the number of litters or the litter size in Pb-exposed dams that began
dosing prior to conception.
There was only a single recent epidemiologic study evaluating the association between Pb
exposure and morphology or histology of female sex organs (ovaries, uterus, fallopian tubes/oviducts,
cervix, vagina, and/or mammary glands) and no recent PECOS-relevant toxicological studies. The results
from the single epidemiologic study reported null associations between blood Pb and rate of uterine
fibroids and uterine fibroid volume, but women with uterine fibroids had higher geometric mean BLLs
than women without fibroids (1.43 (ig/dL versus 1.35 (ig/dL, respectively).
In summary, the collective body of evidence is sufficient to conclude that there is likely to be
a causal relationship between Pb exposure and female reproductive function. The strongest line of
evidence is from recent epidemiologic studies examining the relationship between Pb exposure and
effects on hormone levels and menstrual/estrous cyclicity (Table 8-1). Positive associations from a
longitudinal cohort between bone Pb, a biomarker of cumulative Pb exposure, and both earlier age at
menopause and risk of early menopause were supported by results from a cross-sectional NHANES study
of concurrent exposure of blood Pb with earlier age at menopause. Additionally, recent epidemiologic
studies found consistent positive associations between blood Pb and FSH and LH in women who were
post-menopausal. While these studies are limited by their cross-sectional study design, the studies were
conducted in well-established population-based surveys. These studies considered a range of confounders,
even co-exposure to other metals, but not all studies adjusted for some potential important confounders
such as age at menarche, pregnancy history, oral contraceptive use, and female hormone use, such as IVF
or hormone therapy. While there were no recent PECOS-relevant toxicological studies that examined the
effects of Pb on hormone levels in females or menstrual or estrous cyclicity, previous toxicological
evidence suggests that Pb may disrupt reproductive hormones and menstrual and estrous cyclicity in
females. Two toxicological studies in rats reported disruptions in estrous cyclicity, and two toxicological
studies based in non-human primates reported alterations to different menstrual cycle aspects (e.g., length
of cycle, length of menstruation) and reproductive hormone levels. Additional rodent studies reported
effects of Pb on circulating reproductive hormone levels, including sex steroid hormones (progesterone,
testosterone, and E2) and gonadotropin hormones (LH and FSH).
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8.8.4
Summary of Effects on Male Reproductive Function
The 2013 Pb ISA concluded that there was sufficient evidence to support a causal relationship
between Pb exposures and male reproductive function. This determination was based on toxicological
evidence of sperm/semen production, quality, and function with supporting evidence in the epidemiologic
studies, in addition to evidence supportive for a mode of action. Previous toxicological studies with
relevant Pb exposure routes reported effects on rodent sperm quality and sperm production rate (BLL
range: 34-37 (ig/dL) (Sokol and Berman. 1991; Sokol et al.. 1985). sperm DNA damage (BLL of 19 and
22 (ig/dL) (Nava-Hernandez et al.. 2009). and histological or ultrastructural damage to the male
reproductive organs in studies from rodents (BLL of 5.1 (ig/dL) (El Shafai et al.. 2011) and non-human
primates (BLL of 43 (ig/dL) (Cullen et al.. 1993). These effects were found in animals exposed to Pb
during peripuberty or adulthood for 1 week to 3 months. The toxicological studies reported that Pb
exposure decreased reproductive organ weight and caused histological changes in the testes and germ
cells. Subfecundity (decreased number of pups born/litter) was reported in unexposed females mated to
Pb-exposed males. Also, sperm from Pb-exposed rats (BLLs: 33 to 46 (ig/dL) used for IVF of eggs
harvested from unexposed females yielded lower rates of fertilization (Sokol et al.. 1994). Supporting
evidence was provided by decrements in sperm quality from rabbits administered Pb subcutaneously
(BLLs of 25 (ig/dL) (Moorman et al.. 1998).
The 2013 Pb ISA reported detrimental effects of Pb on sperm observed in epidemiologic studies
with concurrent BLLs of 25 (ig/dL and greater among men occupationally exposed (Hsu et al.. 2009;
Kasperczvk et al.. 2008; Naha and Manna. 2007; Naha and Chowdhurv. 2006). Findings of these
epidemiologic studies are limited due to these high exposure levels among the occupational cohorts and
the lack of consideration for potential confounding factors, including occupational exposures other than
Pb. Studies among men with lower Pb levels were limited to infertility clinic studies, which may produce
a biased sample and findings that lack generalizability. However, a well-conducted epidemiologic study
that enrolled men going to a clinic for either infertility issues or to make a semen donation and controlled
for other metals and smoking reported a positive association of blood Pb with various detrimental effects
in sperm (Telisman et al.. 2007). The median concurrent BLL in this study was 4.92 (ig/dL (range: 1.13-
14.91). A similar study also reported possible associations between concurrent blood Pb and various
semen parameters, but the results were extremely imprecise (large confidence intervals [CIs]), making it
difficult to draw conclusions (Meeker et al.. 2008).
The epidemiologic and toxicological studies in the 2013 Pb ISA reported inconsistent results
regarding hormone aberrations associated with Pb exposure. Due to the complexity of the reproductive
system, uncertainty exists as to whether Pb exerts its toxic effects on the reproductive system by affecting
the responsiveness of the hypothalamic-pituitary-gonadal axis by suppressing circulating hormone levels
or by some other pathway. Inconsistent findings were also apparent among epidemiologic studies of
fertility among men.
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Toxicological studies from the 2013 Pb ISA suggested that oxidative stress was a major
contributor to the effects of Pb exposure on the male reproductive system, providing mode of action
support. The effects of ROS may involve interference with cellular defense systems leading to increased
lipid peroxidation and free radical attack on lipids, proteins, and DNA. Several studies showed that Pb
induced germ cell injury (as evidenced by aberrant germ cell structure and function) and increased
generation of ROS within the male sex organs. Co-administration of Pb with various antioxidant
compounds either eliminated Pb-induced injury or greatly attenuated its effects. In addition, many studies
that observed increased oxidative stress also observed increased apoptosis, which is likely a critical
underlying mechanism in Pb-induced germ cell dysfunction.
Recent epidemiologic and toxicological studies examined Pb exposure and male reproductive
function, including sperm/semen production, quality, and function; hormone levels; fertility; and
morphology and histology of male sex organs. Among the studies that evaluated the relationship of Pb
exposure and sperm/semen production and quality, there was consistent evidence of effects when the
exposure metric was blood Pb. In the recent epidemiologic studies, there were consistent associations of
decreased sperm/semen production and quality with increased blood Pb, but there were inconsistent
associations when Pb was measured in seminal fluid or seminal plasma. The majority of the
epidemiologic studies that evaluated the associations of Pb and sperm/semen production and quality were
cross-sectional studies conducted in males attending fertility clinics, limiting the generalizability of the
results. The studies were further limited by concurrent measurement of exposure and outcome, different
biomarkers of Pb, different seminal parameters, exposure circumstances (historical exposure, magnitude,
duration, timing, and frequency), and small sample sizes. Despite these limitations, it is important to note
that a wide variety of potential confounders were considered, including controlling for hormone levels.
The recent toxicological studies support the findings from the epidemiologic studies. Among the recent
toxicological studies, the majority reported that Pb exposure negatively impacted sperm/semen production
and quality, although these studies were limited to a single species, and no recent toxicological studies
reported on the effects of Pb on sperm or semen parameters in any other laboratory animal species.
There were a limited number of recent epidemiologic and toxicological studies that examined the
relationship between Pb and hormones in males. While recent epidemiologic and toxicological studies
reported changes in hormone levels among males, the direction of the observed relationships differed
across disciplines. Specifically, the epidemiologic studies reported Pb-associated increases in
testosterone, whereas the toxicological studies reported a reduction in testosterone following exposure to
Pb. Additionally, recent cross-sectional epidemiologic studies reported inconsistent associations between
blood Pb and other sex hormones. One study reported positive associations between blood Pb and SHBG,
FSH, and LH, as well as positive trends among quartiles of blood Pb, suggestive of a potential exposure-
response relationship. In contrast, an NHANES study reported null associations between blood Pb and
E2, fE2, androstenedione glucuronide, and SHBG. Recent toxicological evidence regarding the effects of
Pb on male sex hormones is limited to a single study that reported that exposure through the dam's milk
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from birth to weaning (PND 21) in CD-I mice was sufficient to reduce testosterone in the serum at
weaning and on PND 70.
The recent epidemiologic and toxicological studies of Pb exposure and male fertility were
limited. In the recent epidemiologic studies, male fertility was measured by IVF outcomes. There were
inconsistent associations with Pb exposure and male fertility, with one study reporting blood Pb was
associated with longer time to pregnancy, but another reported a positive association between Pb in
seminal plasma and the possibility of obtaining a viable embryo. Differences in Pb biomarkers and
difference in outcomes might explain the inconsistencies of the associations among these studies. The
males in these studies were also recruited from fertility clinics, which might have resulted in selection
bias and limits the generalizability of the results. A single recent toxicological study reported that sperm
from Pb-exposed mice had reduced fertilization capacity, resulting in fewer fertilized oocytes in vitro.
There were a limited number of studies of Pb exposure and morphology or histology of male sex
organs. There was only a single recent epidemiologic study that reported decreased TV, shorter anopenile
distance, and shorter anoscrotal distance with maternal serum Pb exposure. Among the recent
toxicological studies, Pb exposure resulted in effects in the morphology or histology of male sex organs.
Alterations in testis weight were inconsistent with some studies reporting Pb-induced decreases in testis
weight and some reporting that testis weight was unaffected by Pb treatment. However, of the studies that
reported on this outcome, only those that dosed prior to weaning reported that Pb treatment reduced testis
weight, suggesting that this outcome may be more sensitive to developmental exposures. Few studies
investigated the effects of Pb on accessory sex organ weight in males, and of the few studies available, no
effects of Pb were reported on weight of the prostate, seminal vesicles, or epididymides. Testicular
histopathology was consistently altered by Pb exposure, often resulting in visible changes to the
seminiferous tubules and surrounding tissue. Toxicological studies also reported Pb-induced changes in
cellular structures in the epididymides.
In summary, the collective body of evidence is sufficient to conclude a causal relationship
exists between Pb exposure and male reproductive function. There is coherent evidence across the
epidemiologic and toxicological studies of detrimental effects of Pb exposure on male reproductive
function (Table 8-1). The strongest evidence of effects of Pb on male reproductive function is seen in the
consistency of the reported effects of Pb on sperm and semen parameters in both toxicological and
epidemiologic studies. However, the recent epidemiologic and toxicological studies suggest that Pb
exposure may also result in alterations in testosterone levels, fertility, and changes in morphology or
histology of male sex organs. Epidemiologic studies consistently report associations between Pb
measured in blood and decreased sperm/semen production and quality, and toxicological studies
consistently report Pb-induced reductions of a variety of semen parameters such as sperm density,
motility, viability, and normal sperm morphology. There are biological plausible pathways through which
Pb exposure may alter sperm/semen production and quality. Specifically, Pb exposure has been shown to
cause oxidative stress, which can damage the supportive somatic cells in the testis (Leydig cells and
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Sertoli cells) as well as damage the sperm cells directly. Supportive somatic cells are responsible for
producing sex steroid hormones and regulating spermatogenesis, and disruption of either of these
functions can impact the quality and quantity of the sperm produced.
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Table 8-1 Summary of evidence contributing to causality determinations for Pb exposure and reproductive
and developmental effects
Rationale for Causality
Determination3
Key Evidence13
Key References'3
Pb Biomarker Levels Associated with
Effects0
Effects on Pregnancy and Birth Outcomes - Likely to be Causal
A few high-quality
epidemiologic studies of Pb
levels and preterm birth
demonstrate associations
Evidence in a single quasi-experimental study and
some high-quality epidemiologic studies
demonstrates associations with preterm birth. There
is uncertainty related to exposure patterns resulting in
likely higher past Pb exposure, especially among
maternal Pb levels.
Bui etal. (2022)
Jelliffe-Pawlowski et al. (2006)
Viaeh etal. (2011)
See Section 8.3
Maternal BLLs: >10 |jg/dL
A few high-quality
epidemiologic studies show
associations with relevant
BLLs, but findings are
overall inconsistent
Inconsistent findings for studies for maternal health
outcomes, prenatal growth, birth defects,
spontaneous abortion and pregnancy loss, and
placental function. There is uncertainty related to
exposure patterns resulting in likely higher past Pb
exposure, especially among maternal Pb levels.
Additional uncertainties regarding biomarker of
exposure (maternal blood, maternal serum, maternal
bone, maternal erythrocytes, cord blood, cord blood
serum, placental tissue) and the critical window of
exposure.
See Section 8.3
Maternal BLLs: 0.32-6.7 pg/dL
Cord blood Pb: 0.37-10.78 pg/dL
Inconsistent toxicological
evidence
Previous studies report reduced BW, but recent
studies report few impacts of Pb on BW, abortion, still
birth, maternal weight gain, birth defects, or placental
weight and histology.
See Section 8.3
Placental weight altered at BLLs as low
as 12.42 pg/dL
Effects on Development - Causal
Delayed Puberty Onset
Consistent associations
with relevant BLLs in high-
quality epidemiologic
studies
Consistent evidence in multiple cross-sectional and
longitudinal epidemiologic studies for females and
males. Most of these studies have large sample sizes
See Section 8.4.2.1
Section 8.4.3.1
and Female Puberty
BLLs: 0.65-6.57 pg/dL
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Rationale for Causality
Determination3
Key Evidence13
Key References'3
Pb Biomarker Levels Associated with
Effects0
and controlled for potential confounding by covariates
such as age and BMI.
Male Puberty
BLLs: 0.66-6.5
Consistent toxicological
evidence with relevant Pb
exposures
Consistent toxicological evidence from multiple
laboratories of delayed male and female puberty
onset with Pb exposure via diet, drinking water, or
oral gavage in rodents
Pine et al. (2006)
lavicoli et al. (2006)
Dumitrescu et al. (2008a)
Ronis et al. (1998a)
Ronis et al. (1998c)
Dearth etal. (2002)
Dearth etal. (2004)
Ronis et al. (1996)
lavicoli et al. (2004)
Sokoletal. (1985)
Markers of pubertal onset reduced in
animals with BLLs as low as 12.7 |jg/dL
Evidence clearly describes
biological plausibility
Toxicological evidence supports hypothalamic-
pituitary-gonadal axis dysfunction and changes in
IGF-1 contributing to Pb-induced delay in puberty
onset.
Pine etal. (2006)
Dearth etal. (2002)
Pine et al. (2006) reported dam BLLs to
be 39.8 |jg/dL at the time of mating;
Dearth et al. (2002) reported dam BLLs
to be 25.4 |jg/dL at weaning
Postnatal Growth
Available epidemiologic
evidence is inconsistent
Multiple studies, mostly cross-sectional, for children
of varying ages have reported inconsistent results for
the association between BLLs and various measures
of growth. There is uncertainty related to exposure
patterns resulting in likely higher past Pb exposure,
especially among maternal Pb levels. Additional
uncertainties regarding biomarkers of exposure
(maternal blood, maternal serum, maternal bone,
maternal erythrocytes, cord blood, cord blood serum,
placental tissue) and the critical window of exposure.
See Section 8.4.1.1
Maternal BLLs: 0.5-10.1 pg/dL
Cord blood Pb: 0.91-3.1 pg/dL
Available toxicological
evidence is inconsistent
There are inconsistent findings in the toxicological
literature on Pb exposure and postnatal growth.
See Section 8.4.1.2
BLLs ranged from 0.0318-29.16 pg/dL
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Rationale for Causality
Determination3
Key Evidence13
Key References'3
Pb Biomarker Levels Associated with
Effects0
Effects on Female Reproductive Function - Likely to be Causal
A few high-quality
epidemiologic studies of Pb
levels and hormones
demonstrate associations
Evidence in some high-quality cross-sectional
epidemiologic studies demonstrates associations with
hormone levels but results are mixed based on the
hormone examined. There is uncertainty related to
exposure patterns resulting in likely higher past Pb
exposure.
Krieg and Feng (2011)
Chen etal. (2016)
Lee etal. (2019)
BLLs: 1.6-4.1 pg/dL
A few high-quality
epidemiologic studies of Pb
levels and menopause
demonstrate associations
Evidence in some high-quality epidemiologic studies
demonstrates associations with menopause. There is
uncertainty related to exposure patterns resulting in
likely higher past Pb exposure.
Mendola etal. (2013)
Eum etal. (2014)
BLLs: 1.21-3.0 pg/dL
Bone Pb
Tibia: 10 pg/g
Patella: 12 pg/g
Available epidemiologic
studies of Pb levels and
fertility are inconsistent
Epidemiologic studies of this association are limited
by the small sample sizes included in those studies.
In addition, most of the study populations were drawn
from women undergoing IVF and/or attending
infertility clinics. There is uncertainty related to
exposure patterns resulting in likely higher past Pb
exposure.
See Section 8.5.2.1
BLLs: 0.50-2.13 pg/dL
Available toxicological
evidence is inconsistent.
Recent toxicological evidence is scarce and reports
no effects of Pb on litter size and number of litters in
exposed dams. Previous evidence reports
inflammation, decreased ovarian antioxidant capacity,
altered ovarian steroidogenesis.
See Section 8.5
See Section 4.8.4 from U.S.
EPA (2013)
BLLs ranged from 7.72-12.61 pg/dL in
dams in recent literature
Effects on Male Reproductive Function - Causal
Sperm/Semen Production, Quality, and Function
High-quality and consistent Decreased sperm counts, decreased sperm
toxicological evidence with
relevant Pb exposures to
rule out chance, bias, and
confounding with
reasonable confidence.
production rate, dose-dependent suppression of
spermatogenesis in rodents with drinking water Pb
exposure.
See Section 4.8.3.1 from U.S.
EPA (2013)
See Section 8.6.1
BLL after adult drinking water exposure
for 30 d: 34 pg/dL
BLL after peripubertal or adult drinking
water exposure for 30 d: 35 and 37 pg/dL.
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Rationale for Causality
Determination3
Key Evidence13
Key References'3
Pb Biomarker Levels Associated with
Effects0
Ultrastructural and histological damage to non-human
primate testis and seminiferous tubules
Maximum BLLs after daily oral Pb
exposure (gelatin capsule) during infancy,
post infancy, or over a lifetime (up to 10
yr): 32 to 36 pg/dL
Histologic damage to rodent seminiferous tubules
including spermatids and developing sperm.
BLL after adult exposure (oral gavage) for
3 mo: 5.31 pg/dL
Ultrastructural abnormalities to rat spermatogenesis.
BLL after i.p. injection for 16 d: 7.4 pg/dL
Direct effects on rodent sperm DNA after drinking
water Pb exposure.
BLL after adult exposure for 13 wk: 19
and 22 pg/dL
Sperm from Pb exposed rats used for IVF of eggs
harvested from unexposed females yielded lower
rates of fertilization.
BLL after adult exposure for 14-60 d: 33-
46 pg/dL
Semen and sperm quality in rabbits with
subcutaneous Pb treatment; ultrastructural damage
to spermatids with i.p. injection of Pb.
BLL after adult exposure for 15 wk: 16-
24 pg/dL
Available epidemiologic
evidence is inconsistent
The few epidemiologic studies examining this See Section 8.6.1.1
outcome generally have small samples sizes and are
drawn from men attending infertility clinics. There is
uncertainty related to exposure patterns resulting in
likely higher past Pb exposure and biomarker of
exposure (blood, semen, seminal plasma, seminal
fluid).
BLLs: 2.18-3.26 pg/dL
Available toxicological
evidence consistently
reports alterations of sperm
and semen parameters
Consistent reductions of sperm with normal See Section 8.6.1
morphology, sperm density, and sperm viability.
BLLs ranged from 5.09-11.8 pg/dL at
time of outcome assessment
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Rationale for Causality
Determination3
Key Evidence13
Hormone Levels
Key References'3
Pb Biomarker Levels Associated with
Effects0
A few high-quality
epidemiologic studies of Pb
levels and hormones
demonstrate associations
Evidence in some high-quality cross-sectional
epidemiologic studies demonstrates associations with
testosterone levels and adult males, but inconsistent
associations with other hormones. A longitudinal
study among male adolescents reported null
associations with hormone levels.
See Section 8.6.2.1
Concurrent BLLs: 1.0-4.4 |jg/dL
Available toxicological
evidence is inconsistent
Evidence for testosterone is inconsistent across
studies and few studies are available for other male
sex hormones.
See Section 8.6.2
See Section 4.8.3.2 from U.S.
EPA (2013)
Recent study reported effects at BLLs of
5.09 and 19.1 pg/dLattime of outcome
assessment
Fertility
Lack of large, well-
conducted epidemiologic
studies but overall
inconsistent evidence
The few epidemiologic studies examining this
outcome generally have small samples sizes and are
drawn from men attending infertility clinics. There is
uncertainty related to exposure patterns resulting in
likely higher past Pb exposure and biomarker of
exposure (blood, semen).
See Section 8.6.3.1
BLLs: 1.03-1.27 pg/dL
Limited toxicological Few toxicological studies investigate male fertility, but See Section 8.6.3 Recent study reported effects at BLLs of
evidence most report reductions in fertility outcomes such as gee section 4 8 3 3 from U S HQ/dL
fertilized oocytes in recent literature and number of (2013)
offspring per litter in previous studies
BLL = blood lead level; BMI = body mass index; BW = birth weight; d = day(s); IGF-1 = insulin-like growth factor 1; IVF = in vitro fertilization; mo = month(s); Pb = lead; wk = week(s);
yr = year(s).
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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8.9
Evidence Inventories - Data Tables to Summarize Study Details
Table 8-2 Epidemiologic studies of exposure to Pb and maternal health outcomes
Reference and
Study Design Study PoPulatlon
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Gestational Diabetes Mellitus
Shapiro et al. (2015) MIREC
n: 1274
Canada
2008-2011
Cohort
Women at least 18 yr of
age during the first
trimester of pregnancy
(6 to <14 wk gestation)
with singleton, live births
Blood
Maternal blood was
measured by ICP-MS
Age at Measurement:
Maternal age during first
trimester of pregnancy
Geometric mean:
Normal glucose: 0.6 |jg/dL
IGT cases: 0.6 |jg/dL
GDM cases: 0.6 |jg/dL
Quartiles (|jg/dL):
Q1
Q2
Q3
Q4
0.2-0.4
0.5-0.6
0.6-0.9
0.9—4.1
Maternal health during
pregnancy: GDM
IGT and GDM were
assessed by chart review
based on the results of a
50-g glucose challenge test
and 75 or 100-g OGTT
Age at outcome:
Maternal age at IGT or
GDM diagnosis during
pregnancy
Logistic regression models OR (95% CI):
were adjusted for maternal GDM vs norma| g|UCOse
age, race, pre-pregnancy
BMI, and education
Q1
Q2
Q3
Q4
Reference
0.8 (0.3, 1.9)
0.6 (0.2, 1.6)
1.1 (0.5, 2.6)
p for trend: 0.87
IGT vs. normal glucose
Q1
Q2
Q3
Q4
Reference
0.8 (0.4, 1.8)
0.6 (0.2, 1.3)
0.9 (0.4, 2.1)
p for trend: 0.62
GDM or IGT vs. normal
glucose
Q1
Q2
Q3
Q4
Reference
0.8 (0.4, 1.5)
0.6 (0.3, 1.1)
1.0 (0.6, 1.8)
p for trend: 0.76
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Soomro et al. (2019)
Poitiers and Nancy
France
February 2003 to
January 2006
Cohort
Etude des Determinants
pre et post natals du
developpement de la
sante de I'Enfant study
n: 623
Pregnant women
between 24 and 28 wk
of gestation
Blood
Maternal blood measured
by EAAS with Zeeman
background correction
Age at Measurement:
Maternal age at 24-28 wk
gestation
Geometric mean0:
1.62 |jg/dL
Median0: 1.7 pg/dL
75th°: 2.2 pg/dL
95thc: 3.8 pg/dL
Maxc: 8.0 pg/dL
Maternal health during
pregnancy: GDM
At 24-28 wk, maternal
blood glucose
concentrations were
measured 1 hr after a 50 g
glucose challenge. The
GDM was diagnosed by
using the OGTT when there
were >2 blood glucose
concentrations greater than
the following cut points:
fasting = 95 mg/dL, at
1 hr = 180 mg/dL, at
2 hr= 155 mg/dL, and at
3 hr = 140 mg/dL
Age at outcome:
Maternal age at 24-28 wk
gestation
Multiple logistic regression
models were adjusted for
maternal smoking,
maternal age, maternal
BMI, maternal education
level, pregnancy-induced
hypertension, and number
of siblings
OR (95% CI):
GDM vs. normal glucose:
1.318 (0.895, 1.94)
IGT vs. normal glucose:
0.853 (0.676, 1.077)
GDM or IGT vs. normal
glucose: 0.86 (0.682, 1.084)
Oauri et al. (2019)
Japan
January 2011 to
March 2014
Cohort
JECS
n: 16,955
Pregnant women from
15 Regional Centers
throughout Japan who
had single pregnancies,
did not have a history of
diabetes, or receive
insulin treatment, and
hypoglycemic agents
during pregnancy; did
not use steroids during
pregnancy
Blood
Maternal blood was
measured by ICP-MS
Age at Measurement:
Maternal age at 22 to 28 wk
of gestation
Geometric mean
non-GDM: 6.05 ng/g
GDM: 6.13 ng/g
Max: 70.9 ng/g
Quartiles (ng/g):
Q1: <5.00
Q2: 5.1-10.0
Maternal health during
pregnancy: GDM
Pregnant women were
diagnosed with GDM if the
results of a 75 g, 2 hr
OGTT exceeded:
fasting = 92 mg/dL
(5.1 mmol/L);
1 hr = 180 mg/dL
(10.0 mmol/L); and
2 hr= 153 mg/dL
(8.5 mmol/L)
Age at outcome:
maternal age at diagnosis
of GDM
Logistic regression models
adjusted for maternal age
at birth, pre-pregnancy
BMI, pregnancy-induced
hypertension, and pack-
years in the nulliparous
models and maternal age
at birth, pre-pregnancy
BMI, history of GDM,
pregnancy-induced
hypertension, and pack-
years in the parous
models; Model 1 was a
multi-pollutant model with
both Cd and Pb; Model 3
was a single pollutant
model of Pb
OR (95% CI):
Model 1:
Nulliparous:
Q1
Reference
Q2
1.22 (0.75,
1.97)
Q3
1.60 (0.72,
3.55)
Q4
2.51 (0.72,
8.72)
Parous:
Q1
Reference
Q2
0.88 (0.65,
1.20)
Q3
0.79 (0.41,
1.41)
Q4
0.31 (0.04,
2.29)
Model 3:
8-80
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Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Q3: 10.1-15.0
Nulliparous:
Q4: >15.1
Q1
Reference
Q2
1.19 (0.74, 1.91)
Q3
1.55 (0.70, 3.42)
Q4
2.42 (0.70, 8.40)
Parous:
Q1
Reference
Q2
0.89 (0.66, 1.20)
Q3
0.75 (0.41, 1.39)
Q4
0.30 (0.04, 2.23)
Wanqetal. (2019)
Taiyuan
China
2012-2016
Case-control
n: 776 cases and 776
controls
Women aged 18 yr or
older with GA of 20 wk
or more and without
mental illness were
eligible for the study.
Women who had
stillbirths or birth
defects, who had
multiple births, who did
not donate blood
samples, or who had
gestational weeks less
than 29 wk were
excluded
Blood
Maternal blood was
measured by ICP-MS
Age at Measurement:
Mean maternal age for
GDM: 31.00 yr
Mean maternal age for non-
GDM: 30.97 yr
Median0: 2.7968 |jg/dL
75thc: 3.5981 pg/dL
Tertiles (pg/dL):
Low: <2.254
Middle: 2.254-3.323
High: >3.323
Maternal health during
pregnancy: GDM
GDM diagnosis was based
on a 75 g OGTT during
gestational weeks 24 and
28. Women who met one or
more of the following
criteria were diagnosed
with GDM: (1) fasting blood
glucose was more than
5.1 mmol/L, (2) 1 hr blood
glucose >10.0 mmol/L, or
(3) 2 hr blood glucose
>8.5 mmol/L
Age at outcome:
maternal age at gestational
weeks 24-28
Logistic regression models
adjusted for pre-
pregnancy BMI,
gestational weight gain,
physical activity, parity,
family history of diabetes,
and month of conception;
the multi-pollutant model
was also adjusted for
nickel, As, Cd, antimony,
thallium, Hg, and Pb
OR (95% CI):
Single Pollutant Pb Model:
Low: Reference
Middle: 1.04 (0.81, 1.35)
High: 1.01 (00.78, 1.30)
p for trend: 0.963
Multi-pollutant Model:
Low: Reference
Middle: 1.06 (0.80, 1.41)
High: 110 (0.80, 1.51)
p for trend: 0.622
Zhou etal. (2021b)
n: 8169
Blood
Maternal health during
pregnancy: GDM
Logistic regression
analyses: Model 1
OR (95% CI)
Model 1:
China
Pregnant women of GA
in the first trimester
(<14 wk) and singleton
Maternal (serum) analyzed
by polarography method
GDM was diagnosed by the
75 g OGTT according to
adjusted for maternal age,
parity, first trimester BMI,
history of spontaneous
T1: Reference
T2: 1.05 (0.90, 1.21)
abortion, history of ectopic
8-81
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Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
January 2017-
December 2018
Cohort
pregnancy with no
diabetes prior to
pregnancy were
recruited from their first
prenatal visit to the
Southern Medical
University Affiliated
Foshan Women and
Children's Hospital.
Age at measurement:
maternal mean age:
30.14 yr
Median20: 2.53 |jg/dL
75thc: 4.00 pg/dL
Tertiles (pg/dL):
T1
T2
T3
<1.96
1.961-3.41
>3.411
the International
Association for Diabetes in
Pregnancy Study Group's
criteria.
Age at outcome: maternal
mean age: 30.14 yr
pregnancy, family history
of diabetes, family history
of hypertension; Model 2
adjusted for Model 1 plus
other five (Mn, copper,
calcium, zinc, and
magnesium) metals
T3: 0.89 (0.76, 1.03)
Model 2:
T1
T2
T3
Reference
1.05 (0.90, 1.22)
0.89 (0.76, 1.04)
Zheng et al. (2021)
Boston,
Massachusetts
United States
1999-2002
Cohort
Project Viva
n: 1311
Pregnant women
participating in Project
Viva were included in
this study; women were
those of singleton
gestation, able to
answer questions in
English, and GA <22 wk
at recruitment.
Blood
Maternal blood
(erythrocyte) measured in
the first trimester measured
by ICP-MS
Age at measurement:
maternal age during first
trimester mean (SD): 32.3
(4.6) yr
Median: 17.6 ng/g
75th: 23.6 ng/g
Maternal health during
pregnancy: gestational
glucose
Glucose tolerance test 26-
28 wk gestation, as
measured by non-fasting
50 g oral glucose challenge
test.
Age at outcome: maternal
age at 26-28 wk gestation
mean (SD): 32.3 (4.6) yr
BKMR models adjusted
for maternal age, self-
identified race/ethnicity,
pre-pregnancy BMI, GDM
in prior pregnancy,
smoking, maternal
education, diabetes status
of biological mother, and
gestational week at blood
collection for metals
measurements
Difference in mid-gestational
glucose concentration
(mg/dL) associated with IQR
changes of Pb exposure, with
all other metals fixed at their
medians (95% credible
interval)15: -0.5 (-1.6, -0.6)
Tatsuta et al. (2022a) JECS
Japan
2011-2014
Cohort
n: 78,964
Women who delivered a
live infant with singleton
pregnancy. Women
were excluded if there
was a missed blood
sample, missed
Blood
Maternal blood measured
by ICP-MS
Age at measurement:
Maternal age at second or
third trimester; non-GDM
Maternal health during
pregnancy: GDM
GDM diagnosed by OGTT
in second or third trimester
overt GDM diagnosed prior
to OGTT was excluded.
Age at outcome: maternal
Logistic regression
adjusted for pre-
pregnancy BMI, age at
blood collection,
smoking/drinking habits
during pregnancy, history
of GDM, history of
delivering a macrosomia,
regional center, fish intake
OR (95% CI):
Q1
Q2
Q3
Q4
Q5
Reference
1.026 (0.872,
0.968 (0.821,
1.007 (0.854,
0.974 (0.824,
1.206)
1.141)
1.187)
1.151)
8-82
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Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
diagnosis of GDM,
missing HbA1c data,
HbA1c >6.5% at <24
gestational weeks, or a
history of type 1 or type
2 diabetes.
mean age 31.0 yr; GDM
mean age 33.3 yr
Median: 5.9 ng/g
95th: 10.6 ng/g
age at second or third
trimester; non-GDM mean
age 31.0 yr; GDM mean
age 33.3 yr
and co-exposure to Cd,
Mn, and Se
Epigenetic Effects During Pregnancy
Sanders et al. (2015)
Mexico City
Mexico
2007-2011
Cohort
PROGRESS birth cohort
n: 60
This study was
conducted on a sub-
cohort of 60 Mexican
women aged 18-40 yr
participating in the
PROGRESS birth cohort
in Mexico City, and who
consented to a cervical
swab during mid-
pregnancy (16-19 wk
gestation) for miRNA,
thereby participating in
the PROGRESS Cervix
Study.
Blood and bone
Maternal blood was
measured with a dynamic
reaction cell ICP-MS.
Maternal bone was
measured with spot-source
109Cd K-XRF instrument
within 1 mo of delivery
Age at Measurement:
Maternal age at exposure
sampling (mean 27.9 yr
with a range of 18-40)
Mean:
Blood: 2.85 pg/dL
Patellad: 4.16
Tibiad: 1.45
Max:
Blood: 9.38 pg/dL
Patella3: 20.90
Tibia3: 19.45
Maternal health during
pregnancy: altered miRNA
expression in the cervix
Cervical cells were
collected in a method
similar to a standard Pap
smear protocol, where a
cotton swab was used to
collect cells from the
endocervix. Total RNAwas
extracted using the Exiqon
miRCURY kit. MiRNAs
were quantified by using a
NanoPhotometer P-300.
MiRNA expression was
assessed using the
NanoString nCounter
system.
Age at outcome:
Maternal age at
assessment (mean 27.9 yr
with a range of 18-40)
Multivariable linear (3 (95% Cl)b, interpreted as %
regression models were expression change
adjusted for maternal age,
education, smoke
Blood, per 10-fold increase in
exposure in the home, and p^.
panty hsa-miR-297: 84.0 (15.7,
192.8)
hsa-miR-188: 48.5 (7.9,
1.04.2)
Bone Pb, per 1 -unit increase
in Pb:
Patella:
hsa-miR-320e: -4.7 (-7.3,
-1.4)
hsa-miR-22-3p: -4.7 (-8.6,
-0.7)
hsa-miR-93-5p: -6.7 (-12.3,
-0.7)
hsa-miR-769-5p: -5.4 (-9.9,
-0.7)
hsa-miR-297: 2.1 (0.0, 4.2)
hsa-miR-425-5p: -6.7
(-12.3, 0.0)
hsa-miR-361-3p: 2.8 (0.0,
5.7)
Tibia:
8-83
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
hsa-miR-575: -4.1 (-6.7,
-1.4)
hsa-miR-4286: -8.6 (-13.5,
-3.4)
hsa-miR-15a-5p: 7.2 (1.4,
14.1)
hsa-miR-142-3p: 5.7 (0.7,
11.7)
hsa-miR-193b-3p: -7.3
(-12.9, -0.7)
hsa-miR-494: -4.1 (-8.0, 0.0)
Sanchez-Guerra et
al. (2019)
Mexico City, Mexico
December 2007-
2011
Cohort
¦July
PROGRESS Study
n: 410 mother-infant
pairs
Participants who were
<20 wk gestation;
maternal age of >18 yr
and without medical
history of heart or kidney
disease) who underwent
clinical examinations at
different hospitals from
Mexican Social Security
System
Blood
Maternal blood (collected at
second and third trimester
and delivery) and umbilical
cord blood were measured
by ICP-QQQ
Age at measurement:
maternal age at
measurement (mean age
27.22 yr)
Mean
second trimester:
3.79 |jg/dL
3.90 |jg/dL
4.16 |jg/dL
3.50 |jg/dL
75th
second trimester:
third trimester:
at delivery:
cord blood:
4.51 |jg/dL
4.73 |jg/dL
5.28 |jg/dL
4.45 |jg/dL
third trimester:
at delivery:
cord blood:
Maternal health during
pregnancy: altered cord
blood mtDNA content
Venous cord blood
measured the relative
mtDNA content through
mitochondrial-to-nuclear
DNA ratio in cord blood
Age at outcome:
Maternal age at delivery
(Mean age 27.22 yr)
Multivariate linear
regression models were
adjusted for sex, mother's
age, mother's BMI, SES,
smoke exposure, PM2.5
levels, GA, platelets and
leucocytes in cord blood,
C-section, PROM,
preeclampsia, and date of
visit
(3 (95% CI)
Maternal blood
Second trimester: 0.017
(0.002, 0.031)
Third trimester: 0.015 (0.00,
0.03)
At delivery: 0.013 (-0.001,
0.027)
Cord blood
At delivery: 0.016(0.001,
0.03)
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Study"Desfg".? Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Other Outcomes Related to Maternal Health During Pregnancy
Kahn etal. (2014)
Mitrovoca and
Pristina,
Kosovo
May 1985 and
December 1986
Cohort
Yugoslavia Prospective
Study of Environmental
Lead Exposure
n: 291
Women in their second
trimester of pregnancy
were invited to
participate in a study of
pregnancy outcomes at
their first prenatal visit to
government clinics
located at the centers of
two towns in Kosovo.
Women with singleton
births, between 18 and
44 wk of gestation, had
no major central nervous
system defects, no
chromosomal
abnormalities, and
residing <10 km from
clinic
Blood
Maternal blood (serum)
collected at mid-pregnancy
(no method reported)
Age at Measurement:
Pristina Mean: 26.6 yr;
Mitrovica Mean: 26.7 yr
Mean:
Pristina: 5.57 |jg/dL;
Mitrovica: 20 |jg/dL
Max:
Pristina: 18.60 pg/dl_;
Mitrovica: 41.30 |jg/dL
Maternal health during
pregnancy: thyroid function
during pregnancy
Maternal thyroid function
during pregnancy was
assessed using 1T4, TSH,
and TPOAb. 1T4 and
TPOAb were measured by
a radioimmunoassay
procedure, and TSH was
measured using an
immunoradiometric assay
procedure.
Age at outcome:
Pristina Mean: 26.6 yr;
Mitrovica Mean: 26.7 yr
Multiple linear regression
analysis: 1T4 models
adjusted for height,
ethnicity, BMI, fetal GA,
maternal education, adults
per room; TSH models:
hemoglobin, ethnicity,
BMI, fetal GA, maternal
age; TPOAb models
(continuous and
dichotomous): ethnicity,
fetal GA, maternal age,
adults per room
(3 (95% Cl)b
1T4: -0.074 (-0.10, -0.046)
TSH: 0.026 (-0.065, 0.12)
TPOAb: 0.31 (0.17, 0.46)
OR (95% Cl)b
TPOAb: 2.41 (1.53, 3.82),
comparing >10 lU/mLvs.
<10 lll/mL
Wells etal. (2011)
Baltimore, MD
United States
November 2004 and
March 2005
Cohort
Baltimore Tracking
Health Related to
Environmental
Exposures Study
n: 285
Singleton births with
cord blood available,
with complete covariates
data
Blood
UCB was measured by
ICP-MS
Age at Measurement:
Maternal age at delivery
(range: 14-43, mean: 26)
Geometric mean:
0.66 |jg/dL
75th: 0.96 pg/dL
Max: 6.47 pg/dL
Maternal health during
pregnancy: blood pressure
in late pregnancy
Hospital personnel
measured maternal blood
pressure at admission for
labor and delivery and
continuously during
hospitalization. Three pairs
of blood pressure
measurements from each
mother were recorded: SBP
and DBP at admission, the
Multivariable linear
regression models were
adjusted for maternal age,
maternal race,
neighborhood median
household income, prima
parity, smoking during
pregnancy, pre-pregnancy
BMI, and anemia
(3 (95% CI), as change in
blood pressure (mmHg)
Admission SBP:
Q1: Referent
Q2 2.89 (-2.16, 7.94)
Q3: 1.05 (-4.04, 6.14)
Q4: 6.87 (1.51, 12.21)
p for trend: 0.033
Admission DBP:
Q1: Referent
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Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
maximum SBP and
Q2
0.00 (-3.95, 3.96)
corresponding DBP, and
Q3
0.81 (-3.17, 4.80)
the minimum SBP and
Q4
corresponding DBP.
4.40 (0.21, 8.59)
p for trend: 0.036
Age at outcome:
maternal age at delivery
Maximum SBP:
(range: 14-43, mean: 26)
Q1
Referent
Q2
2.47 (-3.08, 8.02)
Q3
-1.76 (-7.36, 3.85)
Q4
7.72 (1.83, 13.60)
p for trend: 0.055
Maximum DBP:
Q1
Referent
Q2
3.93 (-2.86, 10.72)
Q3
-0.42 (-7.27, 6.43)
Q4
8.33 (1.14, 15.53)
p for trend: 0.086
Li etal. (2017b)
Shanghai
China
2010
Cohort
N: 1,485
Pregnant women during
late pregnancy (28-36
gestational weeks)
Blood
Maternal blood was
measured by background
corrected GFAAS collected
gestational week 28-36
Age at measurement:
42 yrold
Geometric mean:
3.99 |jg/dL
Max: 14.84 pg/dL
13-
Maternal health during
pregnancy: maternal stress
Maternal life event stress
and emotional stress were
assessed using the LESPW
and SCL-90-R,
respectively.
Generalized additive
models were adjusted for
maternal age, ethnicity,
maternal education, family
monthly income, years
living in Shanghai
Age at outcome:
old
13-42 yr
(3 (95% Cl)b
Log-blood Pb
GSI: 0.01 (-0.05, 0.07)
Depression: 0.03 (-0.05,
0.10)
Anxiety: 0.01 (-0.06, 0.08)
Log-blood Pb <0.41 |jg/dL
GSI: 0.22 (0.05, 0.40)
Depression: 0.34 (0.12, 0.56)
Anxiety: 0.01 (-0.06, 0.08)
Log-blood Pb >0.41 |jg/dL
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Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
GSI: -0.07 (-0.16, 0.01)
Depression: -0.09 (-0.19,
0.02)
Anxiety: -0.08 (-0.18, 0.02)
Osorio-Yanez et al.
(2021)
Mexico
2007-2011
Cohort
PROGRESS
n: 668
Women enrolled during
second trimester of
pregnancy, were >18 yr
of age, lived in Mexico
City for the following 3 yr
Blood and bone
Maternal blood was
measured by ICP-QQQ;
bone Pb measured by
K-XRF and obtained two
estimated for patella and
tibia (one for each leg),
which were measured 26-
55 d postpartum
Age at measurement:
Median (SD): 27(5.5)yr
Median
Blood - 2nd Trimester:
2.80 |jg/dL
Blood - 3rd Trimester:
2.99 |jg/dL
Bone, tibia: 2.84 |jg/g
Bone, patella: 3.49 |jg/g
Max:
Blood:
2nd Trimester: 20.70 |jg/dL
3rd Trimester: 28.25 |jg/dL
Tibia: 30.1 |jg/g
Patella: 43.2 |jg/g
Maternal health during
pregnancy: bone
remodeling
Bone speed of sound
measured at the second
and third trimesters of
pregnancy at the distal
radius and medium
phalange using QUS.
Age at outcome: Median
(SD): 27 (5.5) yr
Linear models adjusted for
maternal age, SES, parity,
BMI, and GA at the time of
Z-score measurement;
linear mixed model
adjusted for maternal age,
SES, parity, BMI, and GA
at the time of QUS
measurement; models
with blood were mutually
adjusted for other (Cd and
As) metals
(3 (95% Cl)b
Bone (radius) QUS Z-score at
2nd Trimester
Blood (|jg/dL): -0.06 (-0.18,
0.07)
Tibia (|jg/g bone mineral):
0.002 (-0.07, 0.07)
Patella (pg/g bone mineral):
-0.08 (-0.15, -0.01)
Bone (radius) QUS Z-score at
3rd Trimester
Blood (|jg/dL): -0.03 (-0.16,
0.10)
Tibia: 0.017 (-0.05, 0.09)
Patella: -0.03 (-0.10, 0.05)
Bone (radius) QUS Z-score
during pregnancy
Blood (|jg/dL): -0.04 (-0.13,
0.04)
Tibia (|jg/g bone mineral):
0.006 (-0.04, 0.06)
Patella (pg/g bone mineral):
-0.06 (-0.10, -0.01)
Kim et al. (2022)
PROTECT
n: 617
Blood
Maternal health during
pregnancy: MMP
Linear mixed effects
models adjusted for
maternal age, education,
B (95% Cl)b as percent
change in MMP per IQR
increase in blood Pb
8-87
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Puerto Rico and
United States
2010
Cohort
Pregnant women in the
first trimester or early
second trimester of
pregnancy that resided
in the Northern Karst
aquifer region, known
for a large number of
Superfund and other
hazardous waste sites.
Maternal blood, collected at
up to two study visits
(median 18- and 26-wk
gestation), was measured
by ICP-MS
Age at measurement:
Mean (SD) age at
enrollment: 26.9 (5.5) yr
Median:
Enrollment: 0.32 ng/mL
Follow up:0.32 ng/mL
75th:
Enrollment: 0.42 ng/mL
Follow up: 0.43 ng/mL
Max:
Enrollment: 2.18 ng/mL
Follow-up: 1.51 ng/mL
Expression levels of MMP1,
MMP2, and MMP9
measured using
customized Luminex assay
from Invitrogen
Age at outcome: Mean
(SD) age at enrollment:
26.9 (5.5) yr
exposure to second-hand
tobacco smoke, and pre-
pregnancy BMI
MMP1
MMP2
MMP9
23.6 (12.9, 35.2)
5.89 (2.23, 9.67)
-3.31 (-8.12, 1.75)
Females:
MMP1
MMP2
MMP9
Males:
MMP1
MMP2
MMP9
16.3 (5.74, 28.0)
5.48 (1.50, 9.62)
-1.89 (-7.35, 3.90)
10.5 (1.15, 20.6)
2.24 (-1.25, 5.86)
-5.14 (-9.85, -0.17)
Gaiewska et al.
(2021)
Poland
2018-2020
Case-control
n: 146 (66 with
preeclampsia)
Healthy pregnant
women and healthy non-
pregnant women visiting
the Independent Public
Clinical Hospital No 4 in
Lublin for a stay in the
hospital or routine
testing.
Blood
Maternal blood was
measured by ICP-MS
Age at measurement:
Mean: 29.16 yr
Median: 28 yr
Range: 18-47 yr
All Participants:
Mean (SD): 2.63
(1.34) |jg/dL
Median: 2.6 |jg/dL
Preeclampsia Participants:
Maternal health during
pregnancy: preeclampsia
Diagnosis of preeclampsia
was based on the definition
from the American College
of Obstetrics and
Gynecologists.
Age at outcome:
Mean: 29.16 yr
Median: 28 yr
Range: 18-47 yr
Logistic regression
adjusted by the pregnant
woman's age, place of
resident (urban/rural), GA,
multiplicity of pregnancy,
and number of previous
pregnancies
OR (95% Cl)b: 2.65 (1.2,
5.86)
-------
Outcome
Confounders
Effect Estimates and 95%
Clsa
Reference and
Study Design
Study Population
Exposure Assessment
Mean (SD) 3.36 (1.23)
Median: 3.49 |jg/dL
Max:
All Participants: 6.1 |jg/dL
Preeclampsia Participants:
6.1 |jg/dL
Wu etal. (2021) n: 2174
China,
Foshan, Guangdong
Province
August 2019-
November 2019
(participants followed
from 8-12 wk of
pregnancy to birth)
Cohort
Pregnant women that
were registered,
checkup, and delivering
in the Foshan
Chancheng Central
Hospital were included
in the study.
Blood
Maternal blood, collected
between 12 and 27 (±6) wk
of pregnancy and before
date of diagnosed
preeclampsia, was
measured by AAS
Age at measurement:
Mean age at delivery (SD):
29.04 (4.25) yr
Median: 3.60 |jg/dL
Quartiles (|jg/dL):
Q1
Q2
Q3
Q4
2.00-2.90
3.00-3.60
3.70-4.40
4.50-7.90
Maternal health during
pregnancy: preeclampsia
Preeclampsia was based
on electronic medical
records. Preeclampsia was
defined as newly diagnosed
hypertension and
proteinuria occurring after
20 wk of gestation.
Hypertension was defined
as systolic >140 mmHg or
DBP >90 mmHg, 2
occasions, 4 hr apart in a
previously normotensive
woman. Proteinuria was
defined as >300 mg/24-hr
urine collections, or
protein/creatinine >0.3, or
dipstick reading >1
Age at outcome: maternal
age after 28 wk gestation
Logistic regression models OR (95% CI)
were adjusted for age at
delivery, pre-pregnancy
BMI, parity, method of
conception (natural
conception, ART
conception), and
education level; logistic
Dose-effect analysis of the
relationship between BLLs
and the risk of preeclampsia
Linear regression modelb:
1.43 (1.17, 1.74)
BLLs <4.2 |jg/dLb: 0.79 (0.50,
1.24)
BLLs >4.2 |jg/dLb: 2.05 (1.50-
2.81)
Preeclampsia:
Continuous modelb: 1.43
(1.17, 1.74)
Q1: Reference
Q2
Q3
Q4
1.48 (0.64, 3.39)
0.85 (0.33, 2.20)
2.38 (1.13, 5.03)
p for trend: 0.02
Mild Preeclampsia:
Continuous modelb: 1.62
(1.27, 2.06)
Q1: Reference
Q2: 2.63 (0.81, 8.63)
Q3: 1.33 (0.35, 5.06)
Q4: 4.26 (1.41, 12.89)
8-89
-------
Outcome
Confounders
Effect Estimates and 95%
Clsa
p for trend: 0.01
Reference and
Study Design
Study Population
Exposure Assessment
Severe Preeclampsia:
Continuous modelb: 1.10
(0.72, 1.68)
Q1: Reference
Q2: 0.69 (0.19, 2.49)
Q3: 0.51 (0.13, 2.05)
Q4: 1.12 (0.38, 3.27)
p for trend: 0.78
Braun et al. (2014) n: 1054
Mexico City,
Mexico
July 2007 and
February 2011
Cohort
Participants for this
study were enrolled from
an ongoing prospective
birth cohort in Mexico
City. Pregnant women
receiving health
insurance and prenatal
care through the
Mexican Social Security
System were invited to
participate in the study.
To be eligible for
participation in the
study, women had to be
<20 wk gestation, >18 yr
old, free of heart or
kidney disease, have
access to a telephone,
plan to reside in Mexico
City for the next 3 yr, not
use steroids (including
glucocorticoids) or anti-
epilepsy drugs, and not
consume alcohol on a
daily basis.
Blood and bone
Maternal blood was
measured by GFAAS
during the second trimester.
Maternal bone was
measured by K-XRF
instrument ~1 mo
postpartum
Age at measurement:
>18 yr old
Mean:
blood: 3.7 |jg/dL
tibia: 2.7 |jg/g
patella: 4.6 |jg/g
Blood Pb Quintiles
Q1
Q2
Q3
Q4
Q5
0—<1.8 |jg/dL
1.8-<2.4 Mg/dL
2.4-<3.4 |jg/dL
3.4-<5.1 pg/dL
>5.1 pg/dL
Maternal health during
pregnancy: hypothalamic-
pituitary-adrenal axis
function measured from
salivary Cortisol
concentrations
Between 14 and 35 wk of
gestation (mean [SD]: 19.7
[2.4] wk), pregnant women
provided five saliva
samples each day over 2
consecutive days during
the week or weekend.
Women were instructed to
provide samples using the
passive drool technique
upon awakening, 45 min
after waking, 4 hr after
waking, 10 hr after waking,
and at bedtime. Saliva
samples were assayed in
the same batch in duplicate
for Cortisol using a
chemiluminescence assay
with sensitivity of
-0.16 ng/ml.
Linear mixed models with
random intercepts for day
and participant were
adjusted for maternal age,
marital status, years of
education, parity, and
smoking status (never,
former, and current), BMI,
and stress or depressive
symptoms
(3 (95% CI), as % difference in
Cortisol area under the curve
nmol-hr
Blood Pb Quintiles
Q1
Reference
Q2
8 (-1, 18)
Q3
9 (0, 19)
Q4
8 (-1, 18)
Q5
CM
CO
CM
Tibia Pb Quintiles:
Q1
Reference
Q2
-5 (-14, 5)
Q3
2 (-8, 13)
Q4
0 (-10, 10)
Q5
00
CD
Patella Pb Quintiles
Q1
Reference
Q2
1 (-8, 12)
Q3
-6 (-14, 4)
Q4
"3"
CM
CM
8-90
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Tibia Pb Quintiles
Q1
Q2
Q3
Q4
Q5
<2 pg/g
2-<4.3 |jg/g
4.3-<6.7 |jg/g
6.7—<11.1 |jg/g
>11.1 |jg/g
Age at outcome:
maternal age at the time of
outcome measurement
Q5: 4 (-6, 16)
Patella Pb Quintiles
Q1: <2 |jg/g
Q2: 2-<4.5 |jg/g
Q3: 4.5-<7.8 |jg/g
Q4: 7.8-<12.7 pg/g
Q5: >12.7-43.2 |jg/g
Ishitsuka et al. (2020)
Japan
January 2011 -
March 204
Cohort
JECS
n: 17,267
Pregnant women from
15 Regional Centers
throughout Japan who
had single pregnancies,
did not have a history of
diabetes, or receive
insulin treatment, and
hypoglycemic agents
during pregnancy; did
not use steroids during
pregnancy
Blood
Maternal blood was
measured by ICP-MS
Age at measurement:
maternal age at 27 wk of
gestation (mean age:
31 ± 5 yr)
Geometric mean:
0.58 |jg/dL
Quintiles (|jg/dL):
Q1
Q2
Q3
Q4
Q5
0.143-0.433
0.444-0.523
0.524-0.616
0.617-0.7533
0.754-6.752
Maternal health during
pregnancy: maternal
depression
Psychological symptoms
during middle or late
pregnancy were assessed
using the K6.
Age at outcome:
maternal age at 27 wk of
gestation (mean age:
31 ± 5 yr)
Multivariable logistic
regression models
adjusted for age, parity,
marital status, education,
employment status,
household income, and
smoking and alcohol
status
OR (95% CI)
Pb per one-unit increase
K6 >13: 1.00 (0.76, 1.32)
K6 >5: 0.98 (0.88, 1.09)
K6 >13:
Q1
Q2
Q3
Q4
Q5
Reference
0.94 (0.69, 1.27)
0.97 (0.71, 1.31)
0.92 (0.68, 1.25)
0.87 (0.64, 1.19)
K6 >5:
Q1
Reference
Q2
1.03 (0.92, 1.16)
Q3
1.07 (0.95, 1.19)
Q4
0.98 (0.87, 1.10)
Q5
1.01 (0.90, 1.13)
8-91
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Christensen et al.
(2016)
Ukraine and
Greenland
2002-2004
Cross-sectional
Climate Change,
Environmental
Contaminants, and
Reproductive Health
n: 117
Women at least 18 yr
old and born in the
country of the study.
Blood
Maternal blood was
measured by ICP-MS
Age at measurement: >18
Mean0: 1.74 |jg/dL
Median0: 1.457 |jg/dL
Tertiles0 (pg/dL):
T1
T2
T3
0.544-1.013
1.013-1.902
1.902-14.088
Maternal health during
pregnancy: AMH
Concentrations of AMH
were assessed by the
Immunotech enzyme
immunoassay
AMH/Mullerian-inhibiting
substance assay from
serum.
Age at outcome: >18
General linear models
were adjusted for GA,
maternal age, research
site, parity, fish intake,
BMI, ever smoker and
pelvic diseases and
infections
(3 (95% Cl)b per one-unit In-
Pb increase: -0.0423
(-0.4989, 0.4144)
Gustin et al. (2021)
Sweden,
Norrbotten county
Enrollment: 2015-
2018, follow-up
through 29
gestational weeks
Cohort
NICE
n: 544
Pregnant women visiting
their local maternity
clinics who were
residents of southern or
eastern Norrbotten
count and planned to
give birth at Sunderby
Hospital. Only first
pregnancies and
singleton births
included. Those with
thyroid dysfunction were
excluded.
Blood
Maternal blood
(erythrocyte) was measured
by ICP-MS
Age at measurement:
Median: 30 yr
Median: 11 |jg/kg
95th: 27 pg/kg
Maternal health during
pregnancy: hormone Levels
(1T4, tT4, 1T3, tT3, TSH,
1T4:tT4, 1T3:tT3, 1T3:1T4)
Plasma samples from
gestational week 29
analyzed via
electrochemiluminescence
immunoassays
Multivariate linear
regression models
adjusted for parity,
maternal education,
maternal pre-pregnancy
smoking
(3 (95% Cl)b
1T4 (pmol/L): 0.014 (-0.21,
0.18)
tT4 (nmol/L): 0.90 (-1.5, 3.3)
1T3 (pmol/L): 0.036 (-0.018,
0.090)
tT3 (nmol/L): 0.038 (-0.015,
0.091)
TSH (mlU/L): -0.023 (-0.13,
0.087)
1T4:tT4: -0.001 (-0.002,
0.001)
1T3:tT3: -0.009 (-0.031,
0.014)
1T3:fT4: 0.004 (-0.003, 0.011)
Corrales Vargas et
al. (2022)
n: 344
Blood
Maternal blood measured
by ICP-MS
Maternal health during
pregnancy: thyroid function
TSH, 1T4, and 1T3
Linear regression models
adjusted for age, GA,
cotinine detection, pre-
pregnancy BMI, and
(3 (95% CI) for % change in
1st measurement of
outcomes per 10% increase
in blood Pb (pg/L) at
enrollment
8-92
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Matina County,
Limon
Coast Rica
2010-2011
Cohort
Age at measurement:
Maternal age at collection
(recruited <33 wk gestation
with 2nd blood sample
10 wk later)
Median: 0.666 [jg/dL
75th: 0.908 pg/dL
90th: 1.211 pg/dL
Max: 3.43 pg/dL
measured in serum using
electrochemiluminescence
severe vomiting during
pregnancy
TSH (mlU/L): -2.3 (-16.15,
11.55)
fT4 (pmol/L): 0.99 (-0.11,
2.09)
1T3 (pmol/L): -0.21 (-0.52,
0.10)
(3 (95% CI) for % change in
2nd measurement of
outcomes per 10% increase
in blood Pb (pg/L) at
enrollment, excluding outliers:
TSH (mlU/L): -0.08 (-0.22,
0.07)
1T4 (pmol/L): 1.96 (0.66, 3.25)
1T3 (pmol/L): 0.24 (-0.13,
0.61)
AAS = atomic absorption spectrometry; AMH = anti-Mullerian hormone; ART = assisted reproductive technology; BKMR = Bayesian kernel machine regression; BMI = body mass index;
Cd = cadmium; CI = confidence interval; d = day(s); DBP = diastolic blood pressure; EAAS = electrothermal atomic absorption spectrometry; fT3 = free triiodothyronine; fT3:fT4 = ratio of
free triiodothyronine to free thyroxine; fT3:tT3 = ratio of free triiodothyronine to total triiodothyronine; fT4 = free thyroxine; fT4:tT4 = ratio of free thyroxine to total thyroxine;
GDM = gestational diabetes mellitus; GFAAS = graphite furnace atomic absorption spectrometry; GSI = Global Severity Index; hr = hour(s); ICP-MS = inductively coupled plasma mass
spectrometry; ICP-QQQ = inductively coupled plasma triple quad; IGT = impaired glucose tolerance; IQR = interquartile range; JECS = Japan Environment and Children's Study;
K6 = Kessler Psychological Distress Scale; K-XRF = K-shell X-ray fluorescence; LESPW = Life Event Scale for Pregnant Women; MIREC = Maternal-Infant Research on Environmental
Chemicals; miRNA = micro RNA; min = minute(s); MMP = matrix metalloproteinases; mo = month(s); mtDNA = mitochondrial DNA; NICE = Nutritional impact on Immunological maturation
during Childhood in relation to the Environment; OGTT = oral glucose tolerance test; OR = odds ratio; PM25 = fine particulate matter; PROGRESS = Programming Research in Obesity,
Growth, Environment and Social Stressors; PROM = premature rupture of membranes; PROTECT = Puerto Rico Test site for Exploring Contamination Threats; Q = quartile;
QUS = quantitative ultrasound; SBP = systolic blood pressure; SCL-90-R = Symptom-Checklist-90-Revised; SD = standard deviation; Se = selenium; SES = socioeconomic status;
TPOAb = thyroid peroxidase antibodies; TSH = thyroid-stimulating hormone; tT3 = total triiodothyronine; tT4 = total thyroxine; UCB = umbilical cord blood; wk = week(s); yr = year(s).
aEffect estimates are standardized to a 1 pg/dL increase in blood Pb or a 10 pg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect estimates are
standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval. Categorical
effect estimates are not standardized.
bEffect estimates unable to be standardized.
°Pb measurements were converted from pg/L to pg/dL.
dNo units provided.
8-93
-------
Table 8-3
Animal toxicological studies of Pb exposure and pregnancy and birth outcomes
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (pg/dL)
Endpoints
Examined
Saleh et al. (2018)
Rat (Sprague-Dawley)
Control (vehicle), F, n = 8
dams
160 ppm Pb, F, n = 8 dams
320 ppm Pb, F, n = 8 dams
GD 1 to 20 Dams were dosed via oral
gavage. Authors report a
significant decrease in brain
weight occurred, indicating
potential overt toxicity.
Dams (GD 20):
5.1 pg/dL for control
27.7 pg/dL for 160 ppm Pb
41.5 pg/dL for 320 ppm Pb
Abortion, Placental
Weight
Saleh et al. (2019)
Rat (Sprague-Dawley)
Control (vehicle), F, n = 8
dams
160 ppm Pb, F, n = 8 dams
320 ppm Pb, F, n = 8 dams
GD 1 to 20 Dams were dosed via oral
gavage. Authors report a
significant decrease in brain
weight occurred, indicating
potential overt toxicity.
Dams (GD 20):
5.26 pg/dL for control
23.9 pg/dL for 160 ppm Pb
42.9 pg/dL for 320 ppm Pb
Placental Weight
Corv-Slechta et al.
(2013)
Mouse (C57BL/6)
Control (untreated), M/F,
n = 16-29 (8-17/8-12) pups
100 ppm Pb, M/F, n = 16-29
(8-17/8-12) pups
GD -61 to Dams were dosed via
PND 365 drinking water starting 2 mo
prior to mating. Offspring
were continued on the same
exposure as their dams until
the end of the experiment at
12 mo of age. Sample sizes
are only available for "Final"
group sizes for males and
females in Table 1.
Dams at weaning (PND 24):
0.22 pg/dL for control
12.12 pg/dL for 100 ppm Pb
BW, Sex Ratio
Schneider et al.
Mouse (C57BL/6)
GD -61 to
Dams were dosed via
Dams at weaning (assumed BW
(2016)
Control (untreated), F,
PND 21
drinking water starting 2 mo
PND 21): 0.22 pg/dL for control
n = NR
prior to mating through
12.61 pg/dL for 100 ppm Pb
lactation (weaning assumed
100 ppm Pb, F, n = NR
to be PND 21).
Pups (PND 5-6): 0.37 pg/dL for
control
10.2 pg/dL for 100 ppm Pb
Dams were also treated to a
non-stress or prenatal stress
8-94
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (pg/dL)
Endpoints
Examined
condition. Only data from
dams in the non-stress
condition were used.
Wanqetal. (2014)
GD 1-20
Rat (Wistar) GD 1-10, or
Control (untreated), F, n = 17 ®D or
dams
0.25% Pb GD 1-10, F,
n = 16 dams
0.25% Pb GD 11-20, F,
n = 15 dams
Dams were dosed via
drinking water during different
windows of pregnancy.
Assumed termination of study
on GD 20.
Dams (assumed GD 20):
0.828 |jg/dL for control
26.29 |jg/dL for 0.25% Pb GD 1-
10
12.4 |jg/dL for 0.25% Pb GD 11-
20
36.02 |jg/dL for 0.25% Pb GD 1-
20
Placenta
Histopathology,
Placental Weight
0.25% Pb GD 1-20, F,
n = 15 dams
Weston etal. (2014)
Rat (Long-Evans)
Dams
Control (untreated), F, n = 20
50 ppm Pb, F, n = 19
Pups
Control (untreated), M/F,
n = 12.4 (7/5.4 average
number of male and female
pups per litter in control)
50 ppm Pb, M/F, n = 7.4
(6.3/1.1 average number of
male and female pups per
litter in Pb non-stress group)
GD -76 to Dams were dosed via
PND 21 drinking water starting 2-
3 mo prior to breeding.
Exposure ended at weaning
(PND 21).
Dams (PND 21):
0.500 |jg/dL for control
7.72 |jg/dL for 50 ppm Pb
Pups (PND 5-6):
0.603 |jg/dL for control males
0.690 |jg/dL for control females
15.7 |jg/dL for 50 ppm Pb males
14.6 |jg/dL for 50 ppm Pb females
BW, Sex Ratio
Rao Barkur and Bairv
(2016)
Rat (Wistar)
GD -30 to GD -
GD Oto GD 21;
PND 1 to
Dams were dosed via
drinking water for varying
amounts of time:
Pregestation Only (1 mo prior
Pups (PND 22):
0.19 |jg/dL for control
Stillborn Pups, BW
8-95
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (pg/dL)
Endpoints
Examined
Control (untreated), F, n = 6
dams
0.2% Pb Pregestation Only,
n = 6 dams
0.2% Pb Gestation Only,
n = 6 dams
0.2% Pb Lactation Only,
n = 6 dams
0.2% Pb Gestation and
Lactation, F, n = 6 dams
PND21; GD 0 to
PND21
to conception), Gestation
Only (21 d), Lactation Only
(21 d), and Gestation and
Lactation (42 d).
3.03 pg/dL for 0.2% Pb in
Pregestation Only group
5.51 pg/dL for 0.2% Pb in
Gestation Only group
26.86 pg/dL for 0.2% Pb in
Lactation Only group
31.59 pg/dL for 0.2% Pb in
Gestation and Lactation group
Barkur and Bairv
(2015)
Rat (Wistar)
Control (untreated), F, n = 6
dams
0.2% Pb Pregestation Only,
F, n = 6 dams
0.2% Pb Gestation Only,
F, n = 6 dams
0.2% Pb Lactation Only, F,
n = 6 dams
GD-30toGD-1; Dams were dosed via
GD 0 to GD 21;
PND 0 to
PND21; GD 0 to
PND 21
drinking water for varying
amounts of time:
Pregestation Only (1 mo prior
to conception), Gestation
Only (21 d), Lactation Only
(21 d), and Gestation and
Lactation (42 d).
Pups (PND 22):
0.18 pg/dL for control
3.02 pg/dL for 0.2% Pb in
Pregestation Only group
5.30 pg/dL for 0.2% Pb Gestation
Only group
26.7 pg/dL for 0.2% Pb in
Lactation Only group
32.0 pg/dL for 0.2% Pb in
Gestation and Lactation group
Stillborn Pups, BW
0.2% Pb Gestation and
Lactation, F, n = 6 dams
Tartaqlione et al.
(2020)
Rat (Wistar)
Control, M/F, n = NR
50 mg/L Pb, M/F, n = NR
GD -28 to Dams were dosed via
PND 23 drinking water starting 4 wk
prior to mating until weaning
(PND 23).
Pups (PND 23):
0.700 pg/dL for 0 mg/L Pb
25.5 pg/dL for 50 mg/L Pb
BW, Sex Ratio
8-96
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (pg/dL)
Endpoints
Examined
Zhao etal. (2021)
Rat (Sprague-Dawley)
Control (untreated), F, n = 6
dams
109 ppm Pb, F, n = 6 dams
GD-14to Dams were dosed via
PND 10 drinking water starting 2 wk
prior to mating and continued
until PND 10.
Pups:
PND 0
0.87 pg/dL for control
48.2 pg/dL for 109 ppm Pb
PND 10
0.87 pg/dL for control
11.5 pg/dL for 109 ppm Pb
PND 21
0.87 pg/dL for control
2.81 pg/dL for 109 ppm Pb
PND 30
0.87 pg/dL for control
1.20 pg/dL for 109 ppm Pb
BW
Barkuretal. (2011)
Rat (Wistar)
Control (untreated), F, n = 6
dams
0.2% Pb GD Oto PND 21,
F, n = 6 dams
GD Oto PND 21
Dams were dosed via
drinking water throughout
gestation until weaning
(PND 21). Only male pups
were examined.
Pups:
PND 22
0.266 pg/dL for control
31.2 pg/dL for 0.2% Pb
PND 120
0.234 pg/dL for control
0.468 pg/dL for 0.2% Pb
BW
Betharia and Maher
(2012)
Rat (Sprague-Dawley)
Control (untreated), M/F,
n = 36-48 (18-24/18-24)
pups
10 pg/mL Pb, M/F, n = 36-
48 (18-24/18-24) pups
GD 0 to PND 20
Dams were dosed via
drinking water throughout
pregnancy until weaning
(PND 20).
Pups:
PND 2
0.188 pg/dL for control
9.03 pg/dL for 10 pg/mL Pb
PND 25:
Stillborn Pups, Sex
Ratio
8-97
-------
Study
Species (Stock/Strain), n,
Sex
Timing of
Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported (pg/dL)
Endpoints
Examined
0.088 pg/dL for control
0.976 pg/dL for 10 |jg/mL Pb
PND60:
0.0244 pg/dL for control
0.0318 pg/dL for 10 |jg/mL Pb
Graham et al. (2011)
Rat (Sprague-Dawley)
Control (vehicle), M/F,
n = 14-16 (7-8/7-8)
1 mg/kg Pb, M/F, n = 14-16
(7-8/7-8)
10 mg/kg Pb, M/F, n = 14-16
(7-8/78)
PND 4 to 28 Offspring were dosed via oral PND 29:
gavage every other day from 0.267 pg/dL for 0 mg/kg
3.27 pg/dL for 1 mg/kg
12.5 pg/dL for 10 mg/kg
PND 4 until PND 28.
Offspring Mortality
Baranowska-Bosiacka Rat (Wistar) GD 1 to PND 21 Dams were dosed via
etal- (2013) Control (untreated), F, n = 3 drinking water throughout
dams pregnancy until weaning
(PND 21).
0.1% Pb, F, n = 3 dams
Pups (PND 28):
0.93 pg/dL for control
6.86 pg/dL for 0.1% Pb
Sex Ratio
Control, M/F, n = 36 (17/19)
pups
0.1% Pb, M/F, n = 36 (18/18)
pups
BLL = blood lead level; BW = birth weight; d = day(s); F = female; GD = gestational day; M = male; mo = month(s); NR = not reported; Pb = lead; PND = postnatal day; wk = week(s).
8-98
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Table 8-4 Epidemiologic studies of Pb exposure and prenatal growth
Reference^and Study stlldy Poplllatlon Exposllre Assessmen, 0llteome Confcundsrs Effec, Estates ,„d 95%
Xie etal. (2013)
Shandong Province
China
September 2010 and
December 2011
Cohort
n: 252
Pregnant women aged
18 yr or older, planning
to deliver at the Binhai
hospital, and more than
3 yr of residence in the
Laizhou Bay; exclusion
criteria included
diagnoses of
gestational or
preexisting diabetes,
hypertension, HIV, or
AIDS; GA <28 wk;
known occupational
exposure to heavy
metals; with history of
participation in an
assisted reproduction
program; difficulties
with communication;
and infants with severe
neonatal illnesses
Blood and cord blood
Maternal blood and UCB
were measured by GFAAS.
Age at Measurement:
at delivery (within 3 d before
delivery)
Mean (SD):
Maternal: 3.53 (1.51) pg/dL
UCB: 2.92 (1.58) pg/dL
Median:
Maternal: 3.20 pg/dL
UCB: 2.52 pg/dL
75th:
Maternal: 4.18 pg/dL
UCB: 3.95 pg/dL
Max:
Maternal: 11.91 pg/dL
UCB: 10.60 pg/dL
Prenatal growth: BW,
BL, HC
BW, BL, and HC were
measured by several
trained midwives within
1 hr after birth
Age at outcome:
birth
Multiple linear regression
models were adjusted for
infant sex, maternal
education, maternal age,
GA, pre-pregnancy BMI,
parity, and weight gain
during pregnancy
(3 (95% Cl)b
Maternal blood:
BW (g): -148.99 (-286.33,
-11.66)
BL (cm): -0.46 (-1.25,
0.34)
HC (cm): -0.37 (-0.78,
0.19)
UCB:
BW (g): -99.33 (-217.33,
20.67)
BL (cm): -0.84 (-1.52,
-0.16)
HC (cm): -0.36 (-0.81,
0.03)
Garcia-Esquinas et al.
(2013)
Madrid
Spain
October 2003 to May
2004
Cohort
BioMadrid Project
n: 112
Father-pregnant
woman-newborn trios
residing in two areas of
the Madrid
Autonomous Region, a
municipal district in the
city of Madrid (urban
area) and a second
zone lying in the
Greater Madrid
Blood and cord blood
Blood collected from both
parents during pregnancy
and UCB was collected at
delivery and measured by
AAS
Age at measurement:
maternal age: >15; birth
Geometric mean (95% Cl)c
Prenatal growth: BW,
length, 1- and 5-min
Apgar scores
Anthropometric data
were measured once,
before breastfeeding
started. Apgar score
was measured on a
scale from 1 to 10, at 1
and 5 min after
delivery. Infants were
evaluated on a scale of
Multivariable linear
regression models were
adjusted for newborn's
sex, GA, maternal age,
maternal cigarette
smoking and sampling
season
(3 (95% Cl)b
UCB:
BW (g): 123 (-37.9, 284)
BL (cm): 0.52 (-0.39, 1.44)
1-min Apgar score: 0.67
(-0.19, 1.16)
5-min Apgar score: 0.29
(-0.04, 0.54)
8-99
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Metropolitan Belt
(metropolitan area);
women were required
to be aged over 15 yr,
to be expecting a single
pregnancy, and to
intend to deliver their
babies at the public
hospital assigned to
them, lived in the study
area for more than a
year, and did not have
a blood transfusion in
the previous year
Maternal blood: 1.98 (1.816,
2.162) |jg/dL
Paternal blood: 3.30 (3.048,
3.564) |jg/dL
UCB: 1.409 (1.277,
1.555) |jg/dL
Median0:
Maternal blood: 1.898 |jg/dL
Paternal blood: 3.324 |jg/dL
UCB: 1.380 pg/dL
75thc:
Maternal blood: 2.721 pg/dL
Paternal blood: 4.321 pg/dL
UCB: 1.911 pg/dL
0 to 2 according to five
categories (skin color,
muscle tone, reflexes,
respiratory effort, and
heart rate), and the
points from each
category added
together to determine
the total score.
Age at outcome:
birth
Govarts et al. (2016) Flemish human
5 provinces of
Flanders,
Belgium
August 2008-July
2009
Cohort
environmental health
survey (FLEHS II)
n: 248
Women with
uncomplicated live-born
singleton pregnancies,
living in Flanders for at
least 10 yr, ability to fill
in a Dutch
questionnaire, and
giving birth in one out
often randomly
selected maternities
Cord blood
UCB was measured by HR-
ICP-MS
Age at Measurement:
birth
Geometric mean0:
0.864 pg/dL
75thc: 1.138 pg/dL
Prenatal growth: BW Linear regression models (3 (95% Cl)b for an increase
BWwas obtained from
the medical records
Age at outcome:
birth
were adjusted for GA,
child's sex, smoking of the
mother during pregnancy,
parity, and maternal pre-
pregnancy BMI
of Z-score of UCB Pb IQR
increase: -37.14 g (-93.64,
19.36)
Tatsuta et al. (2017) Tohoku Study of Child Cord blood
Tohoku
Japan
2000-2003
Development
n: 489
Singleton pregnancy,
Japanese as the first
language, neonates
UCB was measured by ICP-
MS
Prenatal growth: BW Multiple regression models (3 (p-value)b
BWwas obtained from
medical records
were adjusted for GA,
parity, BMI before
pregnancy,
smoking/drinking habits
All infants: -0.011 g
(0.784)
Male infants: 0.023 g
(0.692)
8-100
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Cohort
born at term (36-42 wk Age at Measurement:
of gestation) with BW of birth
more than 2400 g, and
no congenital Median: 1.0 |jg/dL
anomalies or diseases Ma|e jnfants: 1.0 |jg/dL
Female infants: 1.0 |jg/dL
Age at outcome:
birth
during pregnancy, and
fish/seafood intake
Female infants: -0.039 g
(0.513)
95th: 1.7 |jg/dL
Male infants: 1.7 |jg/dL
Female infants: 1.7 |jg/dL
Wanqetal. (2017b)
Shanghai
China
September 2008 and
October 2009
Cohort
n: 1,009 mother-infant
pairs
Singleton pregnant
women who had lived
in Shanghai for at least
2 yr, were aged 18 yr or
older, and were
delivering at the
selected hospitals were
recruited. Pregnant
women were excluded
if they had chronic
diseases before
pregnancy, pregnancy
complications, or a
history of occupational
heavy metal exposure.
Infants who had severe
disorders or congenital
malformations at birth
were also excluded.
Cord blood
UCB measured by GFAAS
Age at Measurement:
birth
Geometric mean: 4.07 |jg/dL
(95% CI: 3.98, 4.17)
Prenatal growth: BW,
BL, HC, and the PI
Neonatal
anthropometry,
including BW, BL, and
HC, was performed by
trained delivery room
staff with standardized
equipment, and the
results were recorded.
PI was calculated
Age at outcome:
birth
Multiple linear regression
models; models for BW,
HC, and PI were adjusted
for maternal age, GA,
maternal BMI before
delivery, parity, sex of
baby, monthly household
income per capita, mode
of delivery; models for BL
were maternal age, GA,
maternal BMI before
delivery, parity, sex of
baby, monthly household
income per capita; all
models for female infants
and BL model for male
infants were adjusted for
maternal age, GA,
maternal BMI before
delivery, parity, monthly
household income per
capita; models for male
infants for BW, HC, and PI
were adjusted for maternal
age, GA, maternal BMI
before delivery, parity,
monthly household income
(3 (95% Cl)b
All Infants
BW (g): 50.68 (-69.53,
170.88)
BL (cm): 0.36 (-0.13, 0.86)
HC (cm): -0.39 cm (-0.80,
0.02)
PI (g/cm3): -0.03 (-0.12,
0.07)
Female Infants
BW (g): -139.15 (-317.89,
39.59)
BL (cm): 0.32 (-0.38, 1.03)
HC (cm): -0.13 (-0.71,
0.44)
PI (g/cm3): -0.16 (-0.30,
-0.02)
Male Infants
BW (g): 206.50 g (46.15,
366.86)
BL (cm): 0.35 (-0.35, 1.05)
HC (cm): -0.65 (-1.24,
-0.06)
8-101
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RefereDCesignnd Study P°P"'ation
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
per capita, mode of
delivery
PI (g/cm3): 0.09 (-0.04,
0.21)
Govarts et al. (2020)
Belgium
FLEHS I: 2002-2004;
FLEHS II: 2008-2009;
FLEHS III: 2013-
2014; 3xG: 2010-
2015
Cohort
FLEHS I, II and III and
a regional birth cohort
(3xG)
n: 1,579 mother-
newborn pairs: FLEHS
I n = 957, II n = 224, III
n = 273, and 3xG
n = 125
Inclusion criteria were
to be able to fill out a
Dutch questionnaire
and to live at least 5 yr
in the selected study
areas (FLEHS I), at
least 10 yr in Flanders
(FLEHS II), at least 5 yr
in Flanders (FLEHS III),
or living in the
recruitment area (3xG).
Live-born singleton
births
Cord blood
UCB measured by HR-ICP-
MS
Age at Measurement:
birth
Median0:
FLEHS I: 1.42 pg/dL
FLEHS II: 0.83 pg/dL
FLEHS III: 0.61 pg/dL
3xG: 0.61 pg/dL
pooled: 0.97 pg/dL
75thc:
FLEHS I: 2.41 pg/dL
FLEHS II: 1.13 pg/dL
FLEHS III: 0.87 pg/dL
3xG: 0.72 pg/dL
pooled: 1.78 pg/dL
Prenatal growth: BW
BW was recorded
shortly after delivery
Age at outcome:
birth
Multiple linear regression
models were adjusted for
other exposures, GA
(linear and quadratic
terms), sex of the
newborn, maternal age at
delivery, maternal pre-
pregnancy BMI, parity,
smoking during pregnancy
and cohort
(3 (95% Cl)b, interpreted as
the change in mean BW
per interquartile fold
change (the fold change of
the 75th percentile over the
25th percentile in
exposure) in In-Pb: 16.98 g
(-13.14, 47.11) per2.94
interquartile fold change in
In-Pb
Lee et al. (2021)
Dhaka Community
Cord blood
Prenatal growth: BW,
Linear models adjusted for
(3 (95% Cl)b, per IQR
Hospital Trust
BL, HC
maternal age, maternal
increase in In-cord blood
Sirajdikhan and Pabna
n: 1088
UCB measured by acid
digestion and ICP-MS
BMI at enrollment, child
sex, GA, household
income, second-hand
smoke, site daily tea
Pb:
Sadar regions
Trained staff measured
Birth Z-scores
Bangladesh
anthropometer at birth;
GA estimated using
<16-wk ultrasound.
BW (g): -0.04 (-0.19, 0.11)
2008-2011
Age at measurement:
(heavy metals) and cord
BL (cm): -0.06 (-0.20,
birth
blood As, Cd, Mn
0.09)
Cohort
Age at outcome: birth
concentrations
HC (cm): 0.08 (-0.06, 0.23)
Geometric mean (Geometric
SD): 3.18 (2.35) pg/dL
Median: 3.07 pg/dL
Untransformed birth size
measurements
8-102
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
75th: 6.04 pg/dL
Max: 83.5 pg/dL
BW (g): -20.68 (-78.43,
37.08)
BL (cm): -0.23 (-0.61,
0.15)
HC (cm): 0.08 (-0.10, 0.25)
Xu et al. (2012)
Guiyu and Xiamen
China
2001-2008
Cohort
n: 531 (n = 432 from
Guiyu and n = 99 from
Xiamen)
Women who gave birth
in Guiyu or non-urban
area of Xiamen
between 2001 and
2008
Cord blood
UCB measured by GFAAS
Age at Measurement:
birth
Median:
Guiyu: 10.78 pg/dL
Xiamen: 2.25 |jg/dL
Max:
Guiyu: 47.46 |jg/dL
Xiamen: 7.22 |jg/dL
Prenatal growth: BW,
LBW rate, IUGR rate,
GA
Obtained from birth
records; LBW was
defined as <2500 g
Age at outcome:
birth
Multiple linear and logistic (3 (95% Cl)b
regression models were Mean BW (g): _91 80i
adjusted for maternal age
and infant sex
-75)
Mean GA (wk): 0.57 (0.51,
0.63)
OR (95% Cl)b
LBW: 1.61 (1.37, 1.90)
IUGR: 2.12 (1.68, 2.69)
Al-Saleh et al. (2014) n: 1,578
Al-Kharj
Saudi Arabia
2005-2006
Cohort
Women aged 16-50 yr
who delivered in Al-
Kharj hospital, Saudi
Arabia
Cord blood
UCB measured by AAS
Age at Measurement:
maternal age 16-50; birth
Mean (SD):
UCB: 2.551 (2.592) pg/dL
Median:
UCB: 2.057 pg/dL
75th:
UCB: 2.689 pg/dL
Max:
UCB: 56.511 pg/dL
Prenatal growth: PI
PI was calculated as
BW (kilograms) divided
by birth height (m)
cubed
Age at outcome:
birth
Logistic regression model
was adjusted for maternal
age, parity, mother's third
trimester BMI, urinary
cotinine, geographical
distribution of current
dwelling, newborn
mother's highest
education, total family
income, and GA
OR (95% Cl)b: 0.766
(0.502, 1.167)
8-103
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Kim et al. (2020)
Guiyu and Haojiang
China
Cross-sectional
e-REACH Study
n: 314
Women 18 yr or older
with a singleton
pregnancy, had lived in
their respective town
for the duration of their
pregnancy, and
consented to
participate in the study.
Women were excluded
if they had a multiple
pregnancy, used
assistive reproductive
technology to become
pregnant, had a history
of psychiatric or thyroid
disorders, or lived
outside of their
respective town for a
cumulative of 3 mo or
more during their
pregnancy
Blood
Maternal blood, collected at
delivery, was measured by
GFAAS
Age at Measurement:
>18 (age at delivery)
Geometric mean:
Guiyu: 6.7 |jg/dL
Haojiang: 3.8 |jg/dL
Max:
Guiyu: 27 |jg/dL
Haojiang: 16 |jg/dL
Prenatal growth: BW,
HC, GA, newborn BMI,
PI
GA was calculated
based on the LMP and
the date of delivery.
Newborn BMI and PI
were calculated using
the recorded BW and
BL.
Age at outcome:
>18 (age at delivery)
Multiple linear and logistic
regression models were
adjusted for maternal age,
maternal education,
maternal occupation,
maternal BMI, gravidity,
ETS, and neonate sex
(3 (95% Cl)b, interpreted as
the difference in BW, HC,
BMI, or PI, per 1 -unit
increase In-Pb maternal
blood
BW (g): 60 (-15, 135)
HC (cm): -0.75 (-1.17,
-0.32)
BMI (kg/m2):
-0.14 (-0.39, 0.11)
PI (kg/m3): -0.62 (-1.13,
-0.11)
OR (95% Cl)b
SGA: 0.69 (0.33, 1.46)
Xu et al. (2022b)
Ushuaia (South,
higher income) and
Salta (North, lower
income)
Argentina
2011-2012
Cross-sectional
EMASAR
n: 696
Women who either
were about to deliver or
had given birth within
the last 48 hr at one of
the two hospitals.
Women had to be
above 18 yr of age.
Blood
Maternal blood measured by
ICP-MS
Age at measurement:
birth
Median0:
Overall: 1.34 pg/dL
Ushuaia: 0.98 |jg/dL
Salta 1.50 pg/dL
Prenatal growth: GA, Linear models adjusted for (3 (95% CI):
BW, BL, HC, LBW
Medical records were
used to obtain
measures at birth.
Age at outcome: birth
maternal age, pre-
pregnancy BMI, parity,
smoking, and education
BW (g): -47.23 (-94.46,
0.004)
BL (cm):
-0.439 (-0.658, -0.219)
HC (cm):
-0.223 (-0.385, -0.061)
GA(wk): 0.18 (0.05, 0.309)
OR (95%CI)
LBW:
T1: Reference
T2: 0.59 (0.10, 3.55)
8-104
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Geometric mean0:
Overall: 1.393 |jg/dL
Ushuaia: 1.01 |jg/dL
Salta 1.58 pg/dL
T3: 0.53 (0.09, 3.16)
75thc:
Overall: 1.851 pg/dL
Ushuaia: 1.30 |jg/dL
Salta: 2.09 |jg/dL
Hu et al. (2015)
Beijing, Lanzhou,
Taiyuan, Xiamen
China
June-August 2011
Cohort
n: 81
Mother-infant pairs that
were enrolled from 4
hospitals in 4 cities in
China
Blood and cord blood
Maternal blood (serum) and
UCB (serum) were
measured by ICP-MS
Age at Measurement:
median maternal age:
28.5 yr (range: 18-44); at
birth
Prenatal growth: BW
BWwas obtained from
the medical delivery
records
Age at outcome:
birth
Multivariate linear
regression models were
adjusted for infant gender,
maternal age, gestational
week, and maternal BMI
(3 (95% Cl)b
Maternal serum Pb: -1.7 g
(-9.1, 5.6)
UCB serum Pb: -1.5 g
(-5.2, 8.2)
Median:
Maternal: 23.1 ng/g
UCB: 22.0 ng/g
75th:
Maternal: 33.2 ng/g
UCB: 33.7 ng/g
Yang etal. (2020)
Wuhan
China
2014-2015
Births at Women and
Children Medical and
Healthcare Center of
Wuhan
n: 734
The participants were
Cord blood
UCB (serum) was measured
by ICP-MS
Age at Measurement:
birth
Prenatal growth: BW
(BWGA Z-score)
Midwives immediately
measured BW after
delivery and was
standardized for
Generalized linear
regression models
adjusted for maternal age,
annual household income
levels, pre-pregnancy BMI,
parity, passive smoking
during pregnancy,
(3 (95% Cl)b, per unit
increase in In-Pb UCB
serum
Continuous: 0.01 (-0.002,
0.05)
Quartiles:
8-105
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Cohort
enrolled at their first
antenatal examination
^gestational 16 wk).
The inclusion criteria
were (1) residence in
Wuhan city; (2) with a
single gestation; (3)
willing to take the
following prenatal care
during pregnancy and
give birth at the study
hospital; (4) willing to
complete
questionnaires and
provide blood samples
from the umbilical cord
at delivery.
Geometric mean: 1.65 |jg/L
Median: 2.71 |jg/L
75th: 4.29 pg/L
Quartilesd
gestational weeks to
construct BWZ.
Age at outcome:
birth
maternal weight gain
during pregnancy, fetal
sex
Q1
Q2
Q3
<25th percentile
25th-50th percentile
50th-75th percentile
Q4: >75th percentile
Q1
Q2
Q3
Q4
Reference
0.11 (-0.09, 0.30)
-0.05 (-0.24, 0.14)
0.05 (-0.14, 0.24)
p for trend: 0.84
Tang et al. (2016)
Shengsi Island,
Hangzhou
China
July 2011 to May 2012
Cohort
n: 103
Eligible pregnant
women included those
planning to deliver at
the only hospital,
without apparent
clinical symptoms,
without any maternal
history of illness, and
no poor habits such as
drug use. Eligible
infants were singleton
births and had no
congenital diseases.
Cord blood
UCB (serum) measured by
ICP-MS
Age at Measurement:
birth
Mean (SD)C: 12.841
(28.646) pg/dL
Median0: 7.620 pg/dL
75thc: 11.580 pg/dL
Tertiles (pg/dL)c:
T1: <5.633
T2: 5.633-9.197
T3: >9.197
Prenatal growth: BW,
length (height), and HC
and GA
All of these infant
anthropometric
measurements were
collected at birth by
professional healthcare
workers. GAwas
obtained using the
reported date of the
LMP and delivery date.
Age at outcome:
birth
Multivariable linear
regression models were
adjusted for maternal BMI,
maternal age, education
level, newborn gender,
number of abortions,
parity, and pregnancy
weight gain
(3 (95% Cl)b
BW (g): -0.019 (-0.045,
0.006)
BL (cm): 0.29 (-0.50,
-0.09)
HC (cm): -0.22 (-0.44,
-0.00)
GA (wk): -0.21 (-0.44,
0.03)
BW, in g:
T1
T2
T3
Reference
-0.15 (-0.41, 0.11)
-0.05 (-0.30, 0.21)
BL, in cm:
T1
T2
T3
Reference
-0.13 (-0.39, 0.13)
-0.15 (-0.40, 0.11)
8-106
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
HC, in cm:
T1: Reference
T2: -0.31 (-0.59, -0.02)
T3: -0.13 (-0.40, 0.14)
GA, in wk:
T1
T2
T3
Reference
-0.23 (-0.50, 0.05)
-0.20 (-0.49, 0.08)
Freire et al. (2019)
Spain
2000-2008
Cohort
Environment and
Childhood (INMA)
Project
n: 327
Pregnant women of
general population
resident in each study
area [Ribera d'Ebre,
Menorca, Granada,
Valencia, Sabadell,
Asturias and Gipuzkoa]
and their children.
Criteria for inclusion of
the mothers were: (i) to
be resident in one of
the study areas, (ii) to
be at least 16 yr old,
(iii) to have a singleton
pregnancy, (iv) to not
have followed any
program of assisted
reproduction, (v) to
wish to deliver in the
reference hospital and
(vi) to have no
Other: Placenta
Placenta (including maternal
and fetal sides as well as
central and peripheral parts)
measured with GFAAS
using AAS with Zeeman
background correction
Age at Measurement:
birth
Median: <6.5 ng/g (LOD)
75th: <6.5 ng/g (LOD)
Prenatal growth: BW,
length, HC, LBW, GA,
and SGA
Neonatal
anthropometric
measurements were
obtained by the
attending midwife or
nurse; GA was
calculated as the
number of weeks from
the self-reported LMP
to the end of
pregnancy; LBW was
defined by a BW of
less than 2500 g at
term, newborns were
defined as SGA when
below the 10th
percentile of the
expected weight
according to the
Spanish BW curve
adjusted for GA and
sex
Linear models or logistic
regression models were
adjusted for adjusted for
cohort (random effect),
newborn sex, and co-
exposure to other metals
(As, Hg, Cd, Mn, Cr); BW
and LBW models were
additionally adjusted for
GA, maternal smoking
during pregnancy,
maternal working during
pregnancy, and pre-
pregnancy BMI; BL
models were additionally
adjusted for GA and
maternal smoking during
pregnancy; HC models
were additionally adjusted
for GA, maternal smoking
during pregnancy, pre-
pregnancy BMI, and
cesarean delivery; GA
models were additionally
adjusted for maternal
education level; SGA
models were additionally
adjusted for father's
(3 (95% Cl)b
BW (g): 54.57 (-70.84,
180.0)
BL (cm): -0.26 (-0.97,
0.44)
HC (cm): -0.10 (-0.57,
0.36)
GA (wk): -0.11 (-0.57,
0.36)
OR (95% Cl)b
LBW: 2.94 (0.38, 28.34)
SGA: 1.69 (0.53, 8.82)
8-107
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Reference and Study
Design
Study Population
Exposure Assessment Outcome
Confounders Effect Esti™*fs and 95%
communication
problems
Age at outcome:
birth
education and maternal
working during pregnancy
Mikelson et al. (2019) n: 374
Chattanooga, TN
United States
Cohort
Singleton births of HIV
and hepatitis negative
mothers over 18 yr of
age, with GA greater
than 34 wk, and infants
with no major
morphological or
chromosomal
abnormalities
Other: Placenta
Placenta tissue measured
by ICP-MS
Age at Measurement:
birth
Mean (SD): 37.97
(270.5) pg/kg
Median: 12.03 pg/kg
75th: 23.23 pg/kg
Max: 5073 pg/kg
Prenatal growth: BW
Obtained at birth
records
Age at outcome:
birth
Multivariable regression
models adjusted maternal
pre-pregnancy BMI,
maternal age, GA, race,
infant sex, and smoking
while pregnant
(3 (95% Cl)b: -58.3 g
(-97.9, -18.8)
(3 (95% Cl)b, as estimated
change in BW from 25th to
75th percentile: -72.7 g
(-122, -23.4)
Bloom etal. (2015)
Michigan (4 counties)
and Texas (12
counties)
United States
2005-2009
Cohort
LIFE
n: 235
Potential participants
were identified, using
fishing license
registries or a
commercially available
direct marketing data
base, from 12 counties
in Texas and four in
Michigan, respectively,
with presumed
exposure to persistent
organic pollutants.
Inclusion criteria
comprised a committed
heterosexual
relationship, women
aged 18-40 yr (men
>18), English or
Spanish speaker, no
use of an injectable
Blood
Maternal and paternal blood,
collected before pregnancy
(baseline), were measured
by ICP-MS
Age at Measurement:
>18, maternal mean age:
29.75 (SD: 3.73) yr and
paternal mean age: 31.52
(SD:4.57) yr
Mean (SD):
Maternal: 0.71 (0.30) pg/dL
Paternal: 1.13 (0.63) pg/dL
Median:
Maternal: 0.66 pg/dL
Paternal: 0.98 pg/dL
Max:
Prenatal growth: GA,
BW, BL, HC, PI, and
secondary sex ratio
Women were followed
until delivery when they
completed and
returned birth
announcements that
captured date and sex
of birth, weight and
length, and HC.
Secondary sex ratio is
the ratio of live male to
female births, reflecting
a male excess.
Age at outcome:
birth
Multiple regression models
for continuous outcomes:
effect of maternal
exposure adjusted for
paternal exposure,
maternal age, difference in
maternal and paternal age,
and maternal and paternal
smoking, income, race,
serum lipids (mg/dL), and
creatinine for urine
(mg/dL); effect of paternal
exposure adjusted for
maternal exposure,
paternal age, difference in
maternal and paternal age,
and maternal and paternal
smoking, income, race,
serum lipids (mg/dL), and
creatinine for urine
(mg/dL)
(3 (95% CI):
GA, in days
Maternal Exposure:
T1: Reference
T2: 0.43 (-0.48, 1.35)
T3: 0.14 (-0.81, 1.09)
p for trend: 0.671
Paternal Exposure:
T1
T2
T3
Reference
0.19 (-0.70, 1.08)
0.61 (-0.31, 1.53)
p for trend: 0.416
BW, in kg
Maternal Exposure:
T1: Reference
T2: 81.80 (-79.94,
2238.55)
8-108
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
contraceptive within
12 mo, and a menstrual
cycle length of 21-42 d.
Maternal: 2.23 [jg/dL
Paternal: 6.43 [jg/dL
Tertiles (|jg/dL):
Maternal Blood Pb
T1: <0.55 (<33rd percentile)
T2: 0.55-0.73 (33rd to 67th
percentile)
T3: >0.73 (>67th percentile)
Paternal Blood Pb
T1: <0.84 (<33rd percentile)
T2: 0.84-1.16 (33rd to 67th
percentile)
T3: >1.16 (>67th percentile)
T3: -34.885 (-197.76,
128.06)
p for trend: 0.202
Paternal Exposure:
T1: Reference
T2: 20.46 (-134.17,
175.09)
T3: 62.91 (-94.73, 220.55)
p for trend: 0.882
BL, in cm
Maternal Exposure:
T1: Reference
T2: 0.43 (-0.48, 1.35)
T3: 0.14 (-0.81, 1.09)
p for trend: 0.671
Paternal Exposure:
T1
T2
T3
Reference
0.19 (-0.70, 1.08)
0.61 (-0.31, 1.53)
p for trend: 0.416
HC, in cm
Maternal Exposure:
T1: Reference
T2: 0.03 (-0.68, 0.74)
T3: -0.33 (-1.07, 0.41)
p for trend: 0.132
Paternal Exposure:
T1
T2
T3
Reference
0.12 (-0.57, 0.80)
-0.03 (-0.72, 0.67)
p for trend: 0.971
8-109
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
PI, in kg/cm3
Maternal Exposure:
T1: Reference
T2: 0.82 (-7.66, 9.31)
T3: -4.26 (-13.16, 4.64)
p for trend: 0.321
Paternal Exposure:
T1
T2
T3
Reference
-0.22 (-8.50, 8.05)
-5.19 (-13.71, 3.33)
p for trend: 0.150
Rabito etal. (2014)
Shelby County,
Tennessee
United States
2008-2011
Cohort
CANDLE study
n: 98
Healthy pregnant
woman between the
ages of 16 and 40 yr,
carrying a single fetus
with the intent to deliver
the fetus, residence
within Shelby County,
Tennessee, and having
the intent to deliver at
one of three area-
based hospitals
Blood and cord blood
Maternal blood and UCB
were measured by ICP-MS
Age at Measurement:
Maternal age at collection
(second or third trimester or
delivery) (median: 29.50 yr);
birth
Median:
Second trimester:
0.43 |jg/dL
Third trimester: 0.43 |jg/dL
At delivery: 0.50 |jg/dL
Cord blood: 0.37 |jg/dL
Geometric mean (SD):
Second trimester: 0.42
(0.20) |jg/dL
Prenatal growth: BW
BWwas obtained from
medical records
Age at outcome:
birth
Linear regression models
were adjusted for gravidity,
marital status, and GA (in
weeks)
(3 (95% Cl)b, per 0.1 -unit
increase in second
trimester maternal blood
Pb: -43.21 g (-88.6, 2.18)
8-110
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Third trimester: 0.45
(0.28) |jg/dL
At delivery: 0.50
(0.35) |jg/dL
Cord blood: 0.37
(0.32) |jg/dL
Max:
Second trimester:
1.22 |jg/dL
Third trimester: 2.10 |jg/dL
At delivery: 2.47 |jg/dL
Cord blood: 1.80 |jg/dL
Shih etal. (2021)
United States
January 2009-
September 2010
Cohort
Initial Vanguard Study
of the National
Children's Study
n: 125 (68 males, 57
females)
Mother-infant pairs
enrolled in the National
Children's Study.
Blood
Maternal blood was
measured using dynamic
reaction cell ICP-MS.
Age at measurement:
Maternal age at 6-32 wk of
gestation
Median:
Overall: 0.34 |jg/dL
Male infants: 0.35 |jg/dL
Female infants: 0.33 |jg/dL
Max:
Overall: 2.86 |jg/dL
Male infants: 2.86 |jg/dL
Female infants: 0.85 |jg/dL
Prenatal growth: BL,
HC, BW, GA, and PI
Birth outcomes
measured during
physical examination of
infants at birth; BL (cm)
and HC (cm) were
measured twice, and
the average of the two
readings was used. For
those without
measures at birth,
medical records were
extracted by the
National Children's
Study. GA and BW
were obtained from
medical records.
Age at outcome: birth
Linear regression models
adjusted for maternal age,
race/ethnicity, education,
income, smoking status
during pregnancy, number
of prior livebirths,
continuous BMI, and infant
sex
(3 (95% Cl)b, as expected
change for birth outcomes
GA (wk)
Overall: -0.558 (-2.297,
1.181)
Males: 1.084 (-0.855,
3.024)
Females: -4.335 (-7.365,
-1.305)
BW (g)
Overall: -403.593
(-916.671, 109.485)
Males: 141.814 (-431.7,
715.329)
Females: -1685.349
(-2581.105, -789.592)
BL (cm)
Overall: -0.343 (-2.92,
2.233)
Males: 2.211 (-0.667,
5.089)
8-111
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Reference^and Study stlldy Poplllatlon Exposllre Assessmen, 0llteome Confcundsrs Effec, Estates ,„d 95%
Females: -6.37 (-10.86,
-1.88)
HC (cm)
Overall: -1.245 (-2.769,
0.279)
Males: 0.292 (-1.397,
1.981)
Females: -4.866 (-7.52,
-2.212)
PI (kg/m3)
Overall: -3.134 (-6.698,
0.429)
Males: -2.377 (-6.486,
1.731)
Females: -4.733 (-11.191,
1.725)
Woods etal. (2017)
Cincinnati, Ohio
United States
2003-2006
Cohort
HOME Study
n: 272
Women were recruited
between 13 and 19 wk
of pregnancy from
prenatal clinics and
were >18 yr old,
<19 wk gestation at the
time of enrollment, and
living in a residence
built before 1987
Blood
Maternal blood was
measured by sensitive and
specific liquid or gas
chromatography mass
spectrometry
Age at measurement:
maternal age at 16-26 wk
gestation
Geometric mean (geometric
SD): 0.7 (1.4) |jg/dL
Median: 0.7 |jg/dL
75th: 0.8 pg/dL
Prenatal growth: BW
BW was abstracted
from birth records
Age at outcome:
birth
Bayesian hierarchical
linear models were
adjusted for maternal race,
age at delivery, infant sex,
maternal education,
tobacco exposure,
household annual income,
employment, maternal
insurance status, marital
status, pre-natal vitamin
use, and maternal BMI;
sensitivity analysis
included GA
Posterior mean (95%
credible interval)15, as the
difference in BW per 10-
fold increase in maternal
blood Pb: -44.8 g (-110,
21.7)
Taylor et al. (2016) ALSPAC Blood Prenatal growth: BW, Multivariable fractional (3 (95% CI)
n: 4,190 HC, crown-to-heel polynomials and modeled
Bristol, UK length adjusted for maternal
8-112
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
April 1991-December
1992
Cohort
All pregnant women in
the former Avon Health
Authority with an
expected delivery date
between April 1,1991,
and December 31,
1992, were eligible for
the study; 14,541
pregnant women were
initially enrolled,
resulting in a cohort of
14,062 live births
Maternal blood was
measured by ICP-MS
Age at measurement:
maternal age at
measurement (median GA
of sampling: 11 wk)
Median: 3.40 |jg/dL
75th: 4.33 pg/dL
Max: 19.41 pg/dL
HC and CHL were
measured by trained
study staff where the
mother gave
permission or if these
data were missing, the
values were extracted
from the medical
records by trained
study staff. BW was
derived from obstetric
data and from central
birth notification data:
where values
disagreed by <100 g
then the lowest value
was accepted; if the
values disagreed by
>100 g then the value
was coded as missing.
Age at outcome:
birth
educational attainment,
smoking, GA (centered at
40 wk), maternal height
and pre-pregnancy weight,
and sex of the infant
BW (g): -9.93 (-20.27,
0.41)
HC (cm): -0.03 (-0.06,
0.00)
CHL (cm): -0.05 (-0.10,
0.00)
Garcia-Esauinas et al.
(2014)
Madrid
Spain
October 2003-May
2004
Cohort
BioMadrid Project
n: 97
Women were required
to be aged over 15 yr,
to be expecting a single
pregnancy, intend to
deliver their babies at
the public hospital
assigned to them, lived
in the study area
(Madrid Autonomous
Region) for more than a
year, and did not have
Blood and cord blood
Blood, from both parents,
and UCB measured by AAS
with a transversely heated
graphite atomizer furnace
assembly and longitudinal
Zeeman-effect background
correction
Age at measurement:
maternal and paternal age
at median gestational week
was 33.9 (IQR 31.6-35.7)
and at birth
Prenatal growth: GA,
BW, BL, AD, or CD
GA, BW, BL, AD, or
CD was collected at
delivery
Age at outcome:
birth
Multivariable linear
regression models were
adjusted for sampling
maternal age, maternal
tobacco smoke, area
(metropolitan/urban), and
in non-stratified models,
newborn's sex
(3 (95% Cl)b, as mean
difference per two-fold
increase in BLL
Maternal blood Pb
BW (g): 62.4 (-73.1, 197.8)
BL (cm): 0.17 (-0.56, 0.91)
AD (cm): 0.31 (-0.52, 1.15)
CD (cm): 0.15 (-0.21, 0.51)
GA (wk): 0.02 (-0.44, 0.47)
Paternal blood Pb
BW (g): -110.8 (-235.6,
6.0)
8-113
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
a blood transfusion in
BL (cm): -0.44 (-1.12,
the previous year
Geometric mean0:
0.23)
Maternal: 1.83 [jg/dL
AD (cm): -0.81 (-1.64,
Paternal: 3.17 [jg/dL
-0.00)
UCB: 0.45 pg/dL
CD (cm): -0.32 (-0.65,
0.00)
GA(wk): -0.17 (-0.59,
0.26)
UCB Pb
BW (g): 80.0 (-36.8, 196.6)
BL (cm): 0.30 (-0.33, 0.93)
AD (cm): 0.56 (-0.12, 1.24)
CD (cm): -0.16 (-0.47,
0.15)
GA(wk): -0.04 (-0.44,
0.35)
Male Infants
Maternal blood Pb
BW (g): 62.6 (-145.2,
270.4)
BL (cm): 0.29 (-0.83, 1.41)
AD (cm): 1.10 (-0.25, 2.45)
CD (cm): -0.16 (-0.47,
0.15)
GA(wk): 0.11 (-0.58, 0.81)
Paternal blood Pb
BW (g): -93.5 (-269.6,
82.5)
BL (cm): 0.13 (-0.81, 1.06)
AD (cm): -0.64 (-1.89,
0.61)
8-114
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
CD (cm): -0.11 (-0.57,
0.35)
GA(wk): 0.06 (-0.53, 0.65)
UCB Pb
BW (g): 80.0 (-66.0, 226.0)
BL (cm): 0.66 (-0.11, 1.44)
AD: 0.76 cm (-0.16, 1.67)
CD: -0.11 cm (-0.37, 0.39)
GA: 0.06 wk (-0.53, 0.65)
Female Infants
Maternal blood Pb
BW (g): 62.2 (-128.0,
252.4)
BL (cm): 0.08 (-0.95, 1.10)
AD (cm): -0.21 (-1.30,
0.88)
CD (cm): -0.05 (-0.55,
0.46)
GA (wk): -0.06 (-0.70,
0.57)
Paternal blood Pb
BW (g): -129.4 (-312.3,
53.4)
BL (cm): -1.06 (-2.03,
-0.08)
AD (cm): -1.94 cm (-2.06,
0.18)
CD (cm): -0.55 (-1.03,
-0.07)
GA (wk): -0.41 (-1.02,
0.21)
8-115
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
UCB Pb
BW (g): 80.0 (-115.7,
275.7)
BL (cm): -0.37 (-1.41,
0.67)
AD (cm): 0.31 (-0.73, 1.35)
CD (cm): -0.47 (-0.98,
0.05)
GA(wk): -0.13 (-0.79,
0.65)
Daniali et al. (2023)
Isfahan, Iran
2019-2020
Cohort
Prospective
Epidemiologic
Research Studies in
Iran - Isfahan Center
n: 263
Pregnant Iranian
women who have lived
in Isfahan for at
least 1 yr, and did not
have any history of
infertility, those
in the first trimester of
pregnancy, and those
who intended to give
birth in hospitals of
Isfahan city. All
participants with major
risks of SGA and IUGR
such
as serious medical
complications
(hypertension or
diabetes or kidney
disease), cerclage until
24 wk of pregnancy,
Blood
Maternal blood was
measured by ICP-MS
Age at measurement:
maternal age at first
trimester (mean maternal
age 29.94 yr)
Geometric mean ± SD:
2.534 ± 0.205 pg/dL
Median: 2.786 pg/dL
25th: 1.741 pg/dL
75th: 4.01 pg/dL
Prenatal growth: BW,
HC, BL
Standardized neonatal
anthropometric
measurements were
obtained by trained
midwives using
calibrated instruments.
Age at outcome: birth
Infant sex, and maternal
age, BMI at enrollment
(12-14 wk gestation),
income, secondhand
smoke exposure, parity,
and education.
B (95%CI)
BW (g): -0.057 (-0.099,
-0.014)
BL (cm): 0.01 (-0.034,
0.054)
HC (cm): -0.036 (-0.076,
0.004)
8-116
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
history of stillbirth or
preterm labor, multiple
pregnancies, or
abnormal sonographic
evidence were
excluded from the
study.
Taylor et al. (2015)
Bristol
UK
April 1991-December
1992
Cohort
ALSPAC
n: 4,285
All pregnant women in
the former Avon Health
Authority with an
expected delivery date
between April 1, 1991,
and December 31,
1992, were eligible for
the study
Blood
Maternal blood was
measured by ICP-MS
Age at measurement:
maternal age at
measurement (median GA
of sampling: 11 wk)
Mean (SD): 3.67
(1.47) |jg/dL
Geometric mean: 3.43 |jg/dL
Median: 3.42 |jg/dL
Max: 19.14 pg/dL
Prenatal growth: BW,
HC, CHL, and LBW
BW, HC, and CHL
were measured by
trained staff or
extracted from medical
records; LBW was
<2500 g
Age at outcome:
birth
Linear regression models
were adjusted for maternal
height, maternal pre-
pregnancy weight,
maternal educational
attainment, parity, number
of cigarettes per day, sex
of baby, GA at delivery or
death; logistic regression
models for LBW were
adjusted for maternal
height, maternal pre-
pregnancy weight,
maternal educational
attainment, parity, number
of cigarettes per day, sex
of baby and GA at delivery
or death
(3 (95% CI)
BW (g): -1.62 (-2.909,
-0.331)
HC (cm): -0.005 (-0.043,
0.033)
CHL (cm): -0.006 (-0.013,
0.001)
OR (95% CI)
LBW: 1.37 (0.86, 2.18)
Hu et al. (2021)
Canada
2008-2011
Cohort
MIREC
n: 1857
Women from the
MIREC cohort who
delivered singleton live
births, had complete
sociodemographic
information, and
provided biological
samples during the first
trimester of the
pregnancy.
Blood
Maternal blood was
measured by ICP-MS
Age at Measurement:
Maternal age during first
trimester
Geometric mean0:
0.62 |jg/dL
Prenatal growth: BW
Infant BW (g)
abstracted from
medical records and
examined continuously.
Age at outcome: birth
Maternal age, race,
education, pre-pregnancy
BMI, smoking status,
parity, infant sex, cubic-
spline GA; multi-pollutant
model was also adjusted
for As, Cd, Hg, and Mn
j3 (95% CI), as two-fold
increase in Pb blood
Single pollutant model:
-82.22 g (-145.46, -18
.97)
Multi-pollutant model:
-75.89 g (-141.24, -10.54)
8-117
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Median0: 0.60 |jg/dL
75thc: 0.85 pg/dL
Goto etal. (2021)
Japan
January 2011-March
2014
Cohort
JECS
n: 16,423
Pregnant women living
in the study area and
understanding ofthe
Japanese language.
Participants were
excluded: if they did not
meet the Pb
measurement quality
control criteria
(n = 2,002); if mothers
who: were lost to
follow-up; had severe
maternal conditions
preceding pregnancy,
such as chronic
hypertension,
pregestational diabetes
or cardiac disease,
during pregnancy; or
had pregnancies
ending in abortions or
stillbirths (n = 1,209); if
infants had
chromosomal or major
congenital anomalies
(n = 263) or multiple
births (n =283)
Blood
Maternal blood was
measured by ICP-MS
Age at measurement:
maternal age at second or
third trimester (mean age at
delivery: 31 ± 5.0 yr)
Mean: 0.69 pg/dL
Median: 0.63 pg/dL
75th: 0.78 pg/dL
Max: 7.4 pg/dL
Prenatal growth: BW,
SGA, and LBW
BWwas the primary
outcome.
Anthropometric data
were measured by
trained delivery room
staff. Gestational
dating was performed
from the first accurate
ultrasound examination
during the first
trimester. SGA was
defined as a BW below
the 10th percentile of
the national BWs
reported in the
Japanese standard
growth chart, which
also considers GA,
infant sex, and
maternal parity. LBW
was defined as a BW
below 2500 g,
regardless of GA.
Age at outcome:
birth
Multivariable linear
regression models were
adjusted for maternal age
at birth, BMI before
pregnancy, weight gain
during pregnancy,
maternal educational
background, a history of
preterm birth, alcohol
consumption during
pregnancy, smoking habit
during pregnancy, and
parity
j3 (95% CI), per 0.1 pg/dL
increase in maternal blood
Pb
BW (g): -54 (-74.5, -33.5)
HC (cm): -0.10 (-0.05,
-0.15)
BL (cm): -0.20 (-0.30,
-0.10)
GA (days): 0.20 d (-0.35,
0.75)
p (95% CI), per doubling
increment in maternal
blood Pb
BW (g): -86.595 (-112.16,
-61.03)
HC (cm): -0.152 (-0.25,
-0.054)
BL (cm): -0.326 (-0.468,
-0.185)
GA (days): 0.087 (-0.566,
0.74)
OR (95% CI), per 0.1 pg/dL
increase in maternal BLL
SGA: 1.34 (1.16, 1.55)
LBW: 1.34 (1.16, 1.55)
OR (95% CI), per doubling
increment in maternal
blood Pb
SGA: 1.952 (1.526, 2.498)
LBW: 1.34 (1.16, 1.55)
8-118
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Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Rodosthenous et al.
(2017)
Mexico City
Mexico
2007-2011
Cohort
PROGRESS
n: 944
Inclusion criteria:
singleton pregnancy,
GA <20 wk, maternal
age of >18 yr,
expectation to live in
Mexico City for the
following 3 yr, and have
access to a telephone;
exclusion criteria:
chronic medical
conditions such as
heart or kidney
disease; use of steroids
or anti-epilepsy drugs;
drug addiction; and
daily consumption of
alcoholic beverages
due to its association
with adverse fetal
outcomes
Blood
Maternal blood measured by
ICP-QQQ
Age at measurement:
maternal age at -20 wk
gestation
Mean (SD): 3.7 (2.7) pg/dL
Quartile Mean (SD) (pg/dL):
Q1: 1.4 (0.3)
Q2: 2.4 (0.2)
Q3: 3.6 (0.5)
Q4: 7.3 (2.8)
Median: 2.8 pg/dL
75th: 4.5 pg/dL
Max: 22.9 pg/dL
Quartiles (pg/dL)
Q1: <1.93
Prenatal growth:
BWGA, SGA
Infants with a BWGA Z-
score <10th percentile
as SGA
Age at outcome:
birth
Q2
Q3
Q4
1.93-2.79
2.80-4.53
>4.53
Multivariable linear
regression models were
adjusted for maternal age,
BMI, SES, hemoglobin
levels, and infant sex
Quantile regression
models were adjusted for
maternal age, BMI, SES,
hemoglobin levels, and
infant sex
Multivariable logistic
regression models were
adjusted for maternal age,
BMI, SES, hemoglobin
levels, and infant sex
p (95% Cl)b, as difference
in BWGA Z-score per log2
increase in maternal BLL:
-0.06 (-0.013, 0.03)
(3 (95% Cl)b, as the BWGA
Z-score per log2 increase in
maternal BLL
QL 0.05: -0.08 (-0.19,
0.03)
QL 0.10: -0.13 (-0.25,
-0.004)
QL 0.15: -0.11 (-0.22,
-0.002)
QL 0.20: -0.12 (-0.20,
-0.03)
QL 0.25: -0.10 (-0.19,
-0.02)
QL 0.30: -0.11 (-0.18,
-0.04)
QL 0.35: -0.04 (-0.12,
0.04)
QL 0.40: -0.06 (-0.14,
0.03)
QL 0.45: -0.05 (-0.13,
0.04)
QL 0.50: -0.07 (-0.16,
0.01)
QL 0.55: -0.07 (-0.16,
0.01)
QL 0.60: -0.07 (-0.15,
0.01)
QL 0.65: -0.04 (-0.12,
0.04)
QL 0.70: -0.04 (-0.12,
0.03)
8-119
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Confounders
Effect Estimates and 95%
Clsa
QL 0.75: -0.01 (-0.08,
0.06)
QL 0.80: -0.02 (-0.1, 0.06)
QL 0.85: -0.06 (-0.16,
0.04)
QL 0.90: -0.06 (-0.16,
0.02)
QL 0.95: -0.02 (-0.13,
0.09)
OR (95% CI) for SGA:
Q1: Reference
Q2: 1.30 (0.79, 2.15)
Q3: 1.15 (0.92, 1.45)
Q4: 1.09 (1.00, 1.18)
p for trend: 0.06
Ashrap et al. (2020)
Puerto Rico
2010-2017
Cohort
PROTECT
n: 731
Participants were
recruited at
approximately
14 ± 2 wk of gestation
at seven prenatal
clinics and hospitals
throughout Northern
Puerto Rico and
followed until birth;
maternal age between
18 and 40 yr; residence
inside of the Northern
Karst aquifer region;
disuse of oral
contraceptives within
the 3 mo prior to
pregnancy; disuse of
IVF to become
Blood
Maternal blood was
measured by ICP-MS
Age at measurement:
18-40 (collection between
18 and 26 wk of gestation)
Geometric mean (SD):
Preterm births: 0.39
(1.6) pg/dL
Term births: 0.32 (1.5) pg/dL
Median:
Preterm births: 0.36 pg/dL
Term births: 0.32 pg/dL
Prenatal growth: GA,
SGA, LGA, BWZ
All the birth outcome
data were extracted
from medical records.
GA was calculated;
BWZ was defined as
the number of SDs by
which a BW is above or
below the mean; SGA
births were defined as
below the 10th
percentile of BWZs;
LGA births were
defined as above the
90th percentile of
BWZs
Logistic regression models
were adjusted for maternal
age, maternal education
level, pre-pregnancy BMI,
and exposure to second-
hand smoking
(3 (95% Cl)b, per change
per IQR increase in
maternal blood In-Pb
GA (days): -1.
-0.5)
Tertilesd:
GA (days):
(-3.1,
T1
T2
T3
Reference
-0.2 (-2.9, 2.4)
-2.9 (-5.5, -0.2)
BWZ:
T1
T2
T3
Reference
-0.12 (-0.32, 0.07)
0.09 (-0.11, 0.29)
8-120
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Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
pregnant; and free of
any major medical or
obstetrical
complications, including
pre-existing diabetes.
Each woman
participated in a total of
up to three study visits
(18 ± 2 wk, 22 ± 2 wk,
and 26 ± 2 wk of
gestation).
Age at outcome:
birth
OR (95% Cl)b, per change
per IQR increase in
maternal blood In-Pb
SGA: 0.91 (0.69, 1.2)
Tertilesd:
SGA
T1: Reference
T2: 1.58 (0.88, 2.83)
T3: 0.62 (0.30, 1.26)
LGA
T1: Reference
T2: 1.13 (0.63, 2.03)
T3: 0.74 (0.40, 1.40)
Thomas et al. (2015)
Canada
2008-2011
Cohort
MIREC Study
n: 1,835
Pregnant women were
recruited in the first
trimester of pregnancy
from 10 study sites
across Canada.
Exclusion criteria
included: inability to
communicate and
consent in either
French or English,
>14 wk gestation at the
time of recruitment,
<18 yr of age,
diagnosed with a fetal
anomaly or a history of
major chronic disease.
Excluded from the
analysis were: 18
women who withdrew
Blood
Maternal blood, collected
during the first and third
trimesters of pregnancy,
was measured by ICP-MS
Age at Measurement:
Maternal age at first and
third trimesters
Median: 0.59 |jg/dL
75th: 0.81 pg/dL
Max: 4.04 pg/dL
Tertiles (pg/dL):
T1
T2
T3
<0.52
0.52-1.04
>1.04
Prenatal growth: SGA
SGA births were
identified as those
weighing less than the
10th percentile for a
reference population
based on the same
completed week of
gestation and infant
sex
Age at outcome:
birth
Log binomial multivariate
regression models
estimated RR and
adjusted for smoking and
parity
RR (95% CI):
T1: Reference
T2: 1.33 (0.88, 1.99)
T3: 1.19 (0.65, 2.18)
RR (95% CI) for GSTP1
A114V
CC
Pb <0.08 pg/dL: Reference
Pb >0.08 pg/dL: 0.90 (0.57,
1.41)
TC + tT
Pb <0.08 pg/dL: Reference
Pb >0.08 pg/dL: 2.25 (0.95,
5.16)
p for interaction: 0.06
8-121
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Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
during the study, 51
women who gave birth
to multiples, 9
stillbirths, 32
spontaneous abortions,
13 therapeutic
abortions, 28 with no
metal exposure data,
and 15 with no infant
sex, weight, or GA
recorded
RR (95% CI) for GSTP1
1105V
AA
Pb <0.08 |jg/dL: Reference
Pb >0.08 |jg/dL: 1.22 (0.69,
2.15)
AG + GG
Pb <0.08 |jg/dL: Reference
Pb >0.08 |jg/dL: 0.95 (0.54,
1.66)
p for interaction: 0.53
RR (95% CI) for GSTOI
A104A
CC
Pb <0.08 |jg/dL: Reference
Pb >0.08 |jg/dL: 0.94 (0.52,
1.69)
CA + AA
Pb <0.08 |jg/dL: Reference
Pb >0.08 |jg/dL: 1.20 (0.70,
2.06)
p for interaction: 0.54
Ashrapet al. (2021)
Puerto Rico
2011-2017
Cohort
PROTECT
n = 682
Participants were
recruited at
approximately
14 ± 2 wk of gestation
at seven prenatal
clinics and hospitals
throughout Northern
Puerto Rico and
Blood
Maternal blood was
measured by ICP-MS
Age at measurement:
18-40 (collection between
18 and 26 wk of gestation)
Prenatal growth: GA,
BWZ, small for
gestation, large for
gestation
Birth outcomes were
extracted from medical
records. Psychosocial
status was evaluated
using four
questionnaires
Linear and logistic
regression models were
adjusted for maternal age,
maternal education, pre-
pregnancy BMI, second-
hard smoke exposure
(3 (95% Cl)b, per IQR
increase in in maternal
blood In-Pb
GA, change in days
Good Psychosocial Status:
-1.9 (-3.2, -0.6)
Poor Psychosocial Status:
-1.3 (-4.0, 1.5)
BWZ, change in Z-score
8-122
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Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
followed until birth;
maternal age between
18 and 40 yr; residence
inside of the Northern
Karst aquifer region;
disuse of oral
contraceptives within
the 3 mo prior to
pregnancy; disuse of
IVF to become
pregnant; and free of
any major medical or
obstetrical
complications, including
pre-existing diabetes.
Each woman
participated in a total of
up to three study visits
(18 ± 2 wk, 22 ± 2 wk,
and 26 ± 2 wk of
gestation)
Geometric mean: 3.1 |jg/dL Age at outcome: birth
Median: 3.1 |jg/dL
75th 4.1 |jg/dL
95th: 6.5 pg/dL
Max: 15.1 pg/dL
Good Psychosocial Status:
0.1 (0.0, 0.2)
Poor Psychosocial Status:
-0.1 (-0.3, 0.2)
OR (95% Cl)b, per IQR
increase in in maternal
blood In-Pb
SGA
Good Psychosocial Status
0.86 (0.65, 1.14)
Poor Psychosocial Status:
1.49 (0.67, 3.33)
LGA
Good Psychosocial Status
0.89 (0.64, 1.23)
Poor Psychosocial Status:
1.10 (0.57, 2.10)
Gustin et al. (2020)
Norrbotten County
Sweden
2015-2018
Cohort
NICE
n: 589
The cohort was
established in the
catchment area of
Sunderby hospital in
Norrbotten county,
Sweden. At the routine
ultrasound in
gestational week 17-
18, parents who were
interested in
participation were given
more information and
an informed consent to
sign at home and send
back. To be included in
the study, families had
Blood
Maternal blood (erythrocyte)
was measured by ICP-MS
Age at measurement:
Maternal age at gestational
week 24-36 (mean: 31 yr,
range 19-45 yr)
Mean: 14 pg/kg
Median: 11 pg/kg
Max: 148 pg/kg
Prenatal growth: BW,
BL, and HC
Information on the
infants' weight (g),
length (cm), and HC
(cm) at birth was
collected from the
hospital records at
Sunderby hospital.
Age at outcome:
birth
Multivariable-adjusted
linear and spline
regression models were
adjusted for maternal age,
early-pregnancy BMI,
parity, education, pre-
pregnancy smoking, pre-
pregnancy snuff or non-
smoking tobacco use, pre-
pregnancy alcohol
consumption, and
marital/cohabitant status;
infant sex and GA at birth
(in days); models were
also mutually adjusted for
other maternal metals (Cd
and Hg)
(3 (95% Cl)b:
BW (g): -13 (-66, 41)
p for interaction with infant
sex: 0.88
BL (cm): -0.080 (-0.31,
0.15)
p for interaction with infant
sex: 0.43
HC (cm):
Less than median: 0.059
(-0.22, 0.34)
p for interaction with infant
sex: 0.84
Greater than median:
-0.24 (-0.53, 0.056)
8-123
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
to be residents in
Norrbotten county and
be able to
communicate in written
and spoken Swedish.
p for interaction with infant
sex: 0.23
Mutually adjusted for other
maternal metals
BW (g): -0.0091 (-0.077,
0.058)
BL (cm): -0.0078 (-0.079,
0.064)
HC (cm):
Less than median: 0.018
(-0.058, 0.094)
Greater than median:
-0.050 (-0.13, 0.0026)
Rahman et al. (2021) Project Viva
n: 1391
Massachusetts
United States
1999-2002
Cohort
Women were recruited
at prenatal care visits at
eight urban and
suburban practices of a
multi-specialty group
practice in eastern
Massachusetts.
Exclusion criteria
included multiple
gestation, inability to
answer questions in
English, GA >22 wk at
recruitment and plans
to move away from the
study area before
delivery.
Blood
Maternal blood (erythrocyte)
was measured by ICP-MS.
Age at measurement:
maternal age at collection
(mean 11.3 ± 2.8 wk
gestation); mean maternal
age
(SD): 32.3 (4.7) yr
Geometric mean: 17.99 ng/g
Median: 17.7 ng/g
75th: 23.6 ng/g
Prenatal growth: BW,
BL, HC, GA
GA from reported last
menstrual period, BW,
BL, and HC from
medical records
Age at outcome: birth
Multivariable linear
regression models were
adjusted for maternal age,
education, pre-pregnancy
BMI, parity, smoking
status, race/ethnicity,
household income, infant
sex, and GA at delivery
(except when GA is an
outcome)
(3 (95% Cl)b, per IQR
(10.1 ng/g) increase:
BW (g):
Full Cohort: -33.9 (-65.3,
-2.5)
Males: -32.5 (-77.4, 12.5)
Females: -34.6 (-77.2,
8.1)
BL (cm):
Full Cohort: -0.10 (-0.29,
-0.09)
Males: -0.08 (-0.35, 0.19)
Females: -0.13 (-0.39,
0.13)
HC (cm)
Full Cohort: -0.07 (-0.17,
0.04)
Males: -0.14 (-0.29, 0.02)
8-124
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Females: 0.00 (-0.15,
0.15)
GA (wk)
Full Cohort: 0.03 (-0.10,
0.16)
Males: 0.12 (-0.07, 0.30)
Females: -0.04 (-0.22,
0.14)
(95% Cl)b, per IQR
(10.1 ng/g) increase, when
As, Cd, Mn, Zn, and Hg
were fixed at the 25th
percentile:
BW (g):
Full Cohort: -0.05 (-0.11,
0.02)
Males: -0.03 (-0.12, 0.06)
Females: -0.06 (-0.14,
0.02)
BL (cm):
Full Cohort: -0.06 (-0.14,
0.03)
Males: -0.04 (-0.16, 0.07)
Females: -0.04 (-0.14,
0.07)
HC (cm)
Full Cohort: -0.08 (-0.18,
0.02)
Males: -0.08 (-0.26, 0.09)
Females: -0.04 (-0.14,
0.07)
8-125
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
GA (wk)
Full Cohort: 0.02 (-0.06,
0.10)
Males: 0.05 (-0.06, 0.16)
Females: -0.01 (-0.12,
0.100)
(3 (95% CI)b, per IQR
(10.1 ng/g) increase, when
As, Cd, Mn, Zn, and Hg
were fixed at the 50th
percentile:
BW (g):
Full Cohort: -0.04 (-0.10,
0.01)
Males: -0.03 (-0.11, 0.05)
Females: -0.05 (-0.13,
0.02)
BL (cm):
Full Cohort: -0.05 (-0.13,
0.04)
Males: -0.04 (-0.15, 0.07)
Females: -0.03 (-0.14,
0.07)
HC (cm)
Full Cohort: -0.06 (-0.15,
0.03)
Males: -0.06 (-0.21, 0.09)
Females: -0.03 (-0.14,
0.07)
GA (wk)
8-126
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Design
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Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Full Cohort: 0.01 (-0.06,
0.08)
Males: 0.04 (-0.06, 0.14)
Females: -0.02 (-0.12,
0.08)
(3 (95% CI)b, per IQR
(10.1 ng/g) increase, when
As, Cd, Mn, Zn, and Hg
were fixed at the 75th
percentile:
BW (g):
Full Cohort: -0.04 (-0.11,
0.02)
Males: -0.03 (-0.12, 0.06)
Females: -0.05 (-0.13,
0.03)
BL (cm):
Full Cohort: -0.04 (-0.12,
0.05)
Males: -0.04 (-0.15, 0.08)
Females: -0.03 (-0.14,
0.08)
HC (cm)
Full Cohort: -0.03 (-0.13,
0.08)
Males: -0.05 (-0.22, 0.13)
Females: -0.03 (-0.14,
0.08)
GA (wk)
Full Cohort: -0.01 (-0.09,
0.07)
8-127
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Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Males: 0.03 (-0.08, 0.14)
Females: -0.03 (-0.14,
0.08)
Wanqetal. (2017a)
China
2009
Cohort
C-ABCS
n: 3,125
Pregnant women with
singleton, live births
Blood
Maternal blood (serum) was
detected by GFAAS
Age at measurement:
maternal age at collection
(first trimester, median:
11 wk) and second trimester
(median: 16 wk) (mean age:
27.5 yr)
Mean:
Overall: 1.50 |jg/dL
First trimester: 1.52 |jg/dL
Second trimester:
1.49 |jg/dL
Median:
Overall: 1.43 pg/dL
First trimester: 1.43 |jg/dL
Second trimester:
1.43 |jg/dL
Max:
Overall: 5.46 |jg/dL
First trimester: 5.16 |jg/dL
Second trimester:
5.46 |jg/dL
Tertiles (|jg/dL):
Low: <1.18
Medium: 1.18-1.70
Prenatal growth: SGA,
BW, BL, HC, and CC
SGA was defined as
live-born infants with
BW below 10th
percentile for the
babies of the same GA
according to a global
reference; BW, BL, HC,
and CC were
measured at birth
Age at outcome:
birth
Multivariate linear and
logistic regression models
were adjusted for pre-
pregnancy BMI, maternal
age, gravidity, monthly
income, parity, and time of
serum collection
(3 (95% Cl)b
Maternal serum during
pregnancy
BW (g): -2.74 (-5.17,
-0.31)
BL (cm): -0.013 (-0.026,
0.001)
HC (cm): -0.008 (-0.019,
0.004)
CC (cm): -0.008 (-0.018,
-0.002)
First trimester maternal
serum
BW: -4.40 g (-8.22, -0.58)
BL: -0.022 cm (-0.048,
0.005)
HC: -0.007 cm (-0.022,
0.007)
CC: -0.015 cm (-0.030,
<0)
Second trimester maternal
serum
BW (g): -1.64 (-4.80,
-0.58)
BL (cm): -0.006 (-0.020,
0.009)
HC (cm): -0.008 (-0.024,
0.008)
CC (cm): -0.002 (-0.016,
-0.011)
OR (95% CI)
8-128
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Outcome
Confounders
Effect Estimates and 95%
Clsa
High: >1.71
SGA
All Infants
Low: Reference
Medium: 1.45 (1.04, 2.02)
High: 1.69 (1.22, 2.34)
Males
Low: Reference
Medium: 1.44 (0.83, 2.50)
High: 1.75 (1.03, 2.99)
Females
Low: Reference
Medium: 1.51 (0.99, 2.31)
High: 1.68 (1.12, 2.54)
First trimester maternal
serum
Low: Reference
Medium: 1.19 (0.65, 2.19)
High: 2.13 (1.24, 3.38)
Second trimester maternal
serum
Low: Reference
Medium: 1.57 (1.05, 2.34)
High: 1.48 (0.98, 2.21)
Cassidv-Bushrow et
al. (2019)
Wayne County, Ml
United States
September 2003 and
December 2007
(December 2011 and
January 2015)
WHEALS
n: 145
Pregnant women were
in their second
trimester or later, were
aged 21-49 yr, and
were living in a
predefined geographic
area in Wayne and
Oakland counties that
Teeth
Teeth, representing second
and third trimester exposure,
measured by LA-ICP-MS
Mean (SD)6:
Second trimester: 0.04
(0.03) |jg/g
Prenatal growth: BWZ Linear regression models (3 (95% Cl)b
and GA
BWZ and GA obtained
from prenatal and birth
records
Age at outcome:
birth
adjusted for batch, tooth
attrition, tooth type, race,
urban, ETS, anemic,
maternal age, and year
house built; the effect of
time is the difference in
effect estimates from the
second and third
trimesters
BWZ
Second trimester: -0.15
(-0.35, 0.05)
Third trimester: -0.06
(-0.24, 0.12)
Effect of Time: -0.31
(-0.90, 0.28)
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Effect Estimates and 95%
Clsa
Cohort
included the city of
Detroit as well as the
suburban areas
immediately
surrounding the city
Third trimester: 0.05
(0.04) |jg/g
Boys
Second trimester: -0.20
(-0.47, 0.07)
Third trimester: -0.04
(-0.31, 0.23)
Girls
Second trimester: -0.12
(-0.39, 0.15)
Third trimester: -0.06
(-0.33, 0.21)
GA (wk)
Second trimester: 0.08
(-0.19, 0.35)
Third trimester: 0.14
(-0.11, 0.39)
Effect of time: -0.22
(-1.08, 0.64)
Boys
Second trimester: 0.08
(-0.41, 0.57)
Third trimester: 0.01
(-0.44, 0.46)
Girls
Second trimester: 0.12
(-0.21, 0.45)
Third trimester: 0.27
(-0.06, 0.60)
Bui et al. (2022) CCG MSA Births where the mother's Prenatal growth: BW, Difference-in-difference (3 (95% Cl)b, as estimated
n' 147 673 live births in residential address was LBW, and SGA models were used to average treatment effect of
North Carolina the CCG MSA' within 4,000 meters of CMS compare birth outcomes in treatment group
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Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
United States
2004-2009
Quasi-experimental
Treatment group n:
1,138; Control group n:
13,398
Exogenous variation in
Pb exposure resulting
from NASCAR's
deleading of racing fuel
in 2007 was used as a
quasi-experiment.
CMS, located in the
CCG, was the only
NASCAR racetrack in
North Carolina that held
races every year during
our sample period.
Races occurred bi-
annually, in October
and May, ensuring that
all full and near full-
term births in the
sample were prenatally
exposed via the mother
to at least one
NASCAR event.
were classified as the
treatment group, while the
control group consists of
births where the mother's
residential address is in the
CCG but is at least 10,000m
from the racetrack centroid
BWwas the newborn's
weight, in grams. LBW
was defined as BW
<2500 g. SGA was
defined as BW below
the tenth percentile for
clinical GA.
Age at outcome: birth
a non-randomized
treatment group before
and after deleading to
those in the control group.
Models were adjusted for
mother's age, education,
race, and smoking
behavior; father's age,
education, and race;
infant's birth order and
sex; as well as proximity to
a TRI facility or airport,
median household income,
and age of housing stock;
a set of census tract,
month, and year indicator
variables were also
included
BW (g)
All births
Any exposure: 102. 5
(45.73, 152.2)
Trimester 1: 418.6 (205.1,
632.1)
Trimester 2: 47.68 (-40.01,
135.4)
Trimester 3: 262 (97.01,
427.1)
Full-term births
Any exposure: 24.08
(-15.14, 63.29)
Trimester 1: 104.7 (-54.65,
264)
Trimester 2: 44.16 (-36.35,
124.7)
Trimester 3: 80.19 (-30.44,
190.8)
LBW
All births
Any exposure: -0.045
(-0.07, -0.019)
Trimester 1: -0.062
(-0.178, 0.054)
Trimester 2: -0.022
(-0.061, 0.017)
Trimester 3: -0.158
(-0.314, -0.001)
Full-term births
Any exposure: 0.001
(-0.014, 0.016)
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Effect Estimates and 95%
Clsa
Trimester 1: 0.05 (-0.038,
0.138)
Trimester 2: -0.035
(-0.054, -0.016)
Trimester 3: -0.006 (-0.07,
0.057)
SGA
All births
Any exposure: -0.04
(-0.064, -0.016)
Trimester 1: -0.042
(-0.242, 0.158)
Trimester 2: -0.058
(-0.118. 0.002)
Trimester 3: -0.038
(-0.122, 0.045)
Full-term births
Any exposure: -0.028
(-0.051, -0.004)
Trimester 1: -0.053
(-0.274, 0.168)
Trimester 2: -0.049
(-0.103, 0.004)
Trimester 3: 0.022 (-0.081,
0.125)
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Confounders
Effect Estimates and 95%
Clsa
AAS = atomic absorption spectrometry; AD = abdominal diameter; ALSPAC = Avon Longitudinal Study of Parents and Children; As = arsenic; BL = birth length; BMI = body mass
index; BW = birth weight; BWGA = birth weight-for-gestational age; BWZ = birth weight Z-score; C-ABCS = China-Anhui Birth Cohort Study; CANDLE = Conditions Affecting
Neurocognitive Development and Learning in Early Childhood; CC = chest circumference; CCG = Charlotte-Concord-Gastonia; Cd = cadmium; CD = cephalic diameter;
CHL = crown-heel length; CMS = Charlotte Motor Speedway; Cr = chromium; d = day(s); EMASAR = Study on the Environment and Reproductive Health; e-REACH = e-waste
Recycling Exposure and Community Health; ETS = environmental tobacco smoke; FLEHS = Flemish Environment and Health Study; GA = gestational age; GFAAS = graphite
furnace atomic absorption spectrometry; HC = head circumference; Hg = mercury; HOME = Health Outcomes and Measures of the Environment; hr = hour(s); HR-ICP-MS = high
resolution inductively coupled plasma mass spectrometry; ICP-MS = inductively coupled plasma mass spectrometry; ICP-QQQ = inductively coupled plasma triple quad;
INMA = Instituto de Nanociencia y Materiales de Aragon; IQR = interquartile range; IUGR = intrauterine growth restriction; IVF = in vitro fertilization; LA-ICP-MS = laser ablation-
inductively coupled plasma-mass spectrometry; LBW = low birth weight; LGA = large for gestational age; LIFE = Longitudinal Investigation of Fertility and the Environment;
LMP = last menstrual period or last missed period; In = natural log; LOD = limit of detection; MIREC = Maternal-Infant Research on Environmental Chemicals; min = minute(s);
Mn = manganese; mo = month(s); MSA = Metropolitan Statistical Area; NICE = Nutritional impact on Immunological maturation during Childhood in relation to the Environment;
OR = odds ratio; PI = Ponderal Index; PROGRESS = Programming Research in Obesity, Growth Environment and Social Stress; PROTECT = Puerto Rico Test site for Exploring
Contamination Threats; QL = lower quartile; RR = relative risk; SD = standard deviation; SES = socioeconomic status; SGA = small for gestational age; TRI = Toxics Release
Inventory; UCB = umbilical cord blood; WHEALS = Wayne County Health, Environment, Allergy and Asthma Longitudinal Study; wk = week(s); yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bEffect estimates unable to be standardized.
°Pb measurements were converted from |jg/L to |jg/dL.
dNo cut points provided for the categorizations.
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Table 8-5
Epidemiologic studies of Pb exposure and preterm birth
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Xu et al. (2012)
Guiyu and Xiamen
China
2001-2008
Cohort
n: 531 (n = 432 from Guiyu
and n = 99 from Xiamen)
Women who gave birth in
Guiyu or non-urban area of
Xiamen between 2001 and
2008
Cord blood
UCB measured by GFAAS
Age at measurement:
birth
Median:
Guiyu: 10.78 pg/dL
Xiamen: 2.25 |jg/dL
Max:
Guiyu: 47.46 |jg/dL
Xiamen: 7.22 |jg/dL
Preterm birth rate Multiple logistic regression
models were adjusted for
Preterm birth was defined maternal age and infant sex
as birth <37 wk gestation
Age at outcome: birth
OR (95% Cl)b: 1.09
(0.93, 1.28)
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Design
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Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Xu et al. (2022b)
Argentina
2011-2012
Cross-sectional
EMASAR
696
n
Women who either were
about to deliver or had given
birth within the last 48 hr at
one of the two hospitals.
Women had to be above
18 yr of age.
Blood
Maternal blood measured
by ICP-MS
Age at measurement:
birth
Median0:
Overall: 1.34 pg/dL
Ushuaia: 0.98 |jg/dL
Salta 1.50 pg/dL
Geometric mean0:
Overall: 1.393 pg/dL
Ushuaia: 1.01 pg/dL
Salta 1.58 pg/dL
75th°:
Overall: 1.851 pg/dL
Ushuaia: 1.30 pg/dL
Salta: 2.09 pg/dL
Preterm birth
Medical records were
used to obtain measures
at birth.
Age at outcome: birth
Logistic models adjusted for
maternal age, pre-pregnancy
BMI, parity, smoking,
education, and LBW
OR (95%CI)
T1
T2
T3
Reference
1.24 (0.35,
1.26 (0.32,
4.40)
5.00)
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Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Freire et al. (2019)
Spain
2000-2008
Cross-sectional
INMA Project
n: 327
Pregnant women of general
population resident in each
study area [Ribera d'Ebre,
Menorca, Granada,
Valencia, Sabadell, Asturias
and Gipuzkoa] and their
children. Criteria for
inclusion of the mothers
were: (1) to be resident in
one of the study areas, (2)
to be at least 16 yr old, (3) to
have a singleton pregnancy,
(4) to not have followed any
program of assisted
reproduction, (5) to wish to
deliver in the reference
hospital and (6) to have no
communication problems
Other: Placenta
Placenta (including
maternal and fetal sides as
well as central and
peripheral parts) measured
by GFAAS with Zeeman
background correction
Age at measurement:
birth
Median: <6.5 ng/g (LOD)
75th: <6.5 ng/g (LOD)
Preterm delivery
Preterm birth was defined
as live birth before 37 wk
of pregnancy,
Age at outcome:
birth
Logistic regression models
were adjusted for cohort
(random effect), newborn sex,
co-exposure to other metals
(As, Hg, Cd, Mn, Cr), and
maternal education level
OR (95% Cl)b:
(0.04, 4.70)
0.40
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Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Yu et al. (2019)
Shanxi Province
China
December 2009-
December 2013
Case-control
n: 528
Women with prenatal
examination at <22
gestational weeks, >18 yr
old, and living in the local
counties for >1 yr
Blood
Maternal blood (serum) was
measured by ICP-MS
Age at measurement:
maternal age at first
trimester (<12 wk gestation)
or second trimester (13-
28 wk)
Mediand
Overall: 0.0482 pg/dL
First trimester:
0.0489 pg/dL
Second trimester:
0.0476 pg/dL
75thd:
Overall: 0.0751 pg/dL
First trimester:
0.0783 pg/dL
Second trimester:
0.0735 pg/dL
Spontaneous preterm
birth
Spontaneous preterm
birth is defined as a live
birth at <37 wk GA without
iatrogenic causes,
including spontaneous
preterm labor with intact gender
membranes and PROM
Age at outcome:
birth
Unconditional logistic
regression models were
adjusted for maternal age, BMI,
education, occupation,
residence, gravidity, parity,
spontaneous abortion history,
folic acid use, drug use,
passive smoking, and child
OR (95% Cl)b
Overall: 1.46 (0.97, 2.18)
First trimester: 1.63
(0.91, 2.91)
Second trimester: 1.27
(0.71, 2.28)
Xu et al. (2022a)
Pingding, Shouyang,
and Taigu Counties
Shanxi Province
China
December 2009-
December 2013
Case-Control
n: 148 (74 cases, 74
controls)
Pregnant women were
recruited if over 18 yr old,
living locally for at least 1 yr,
seeking first prenatal visit at
or before 22 gestational
weeks, and seeking to
manage birth/pregnancy at
Maternal and Child Health
Hospitals of study counties.
Blood
Maternal blood (serum)
measured by ICP-MS
Age at measurement:
maternal age during 4-22
gestational week
Mediand: 0.049 pg/dL
75thd: 0.078 pg/dL
Spontaneous preterm
birth
Information about
spontaneous preterm birth
was collected from
pregnancy health records
at the hospitals
Age at outcome: birth
Unconditional logistic
regression with adjustment for
age, BMI, education,
occupation, residence,
gravidity, parity, spontaneous
abortion history, folic acid use,
medication use, passive
smoking, infant sex, fasting
blood collection, and sampling
time.
OR (95% CI):
Q1
Q2
Q3
Q4
Reference
1.63 (0.53, 5.04)
1.81 (0.60, 5.52)
4.09 (1.31, 12.77)
p for trend: 0.017
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Tsuii etal. (2018)
Japan
January 2011 and
March 2014
Cohort
JECS
n: 14,847
Women who delivered live
birth infant with singleton
pregnancies without missing
exposure or covariate data
Blood
Maternal blood measured
by ICP-MS
Age at measurement:
Maternal age at gestational
weeks 14-39 (mean
maternal age 31.4 yr)
Preterm birth
Preterm births were
divided into early (<34 wk)
and late preterm births
(34 to <37 wk)
Age at outcome:
birth
Multivariable logistic regression
analysis adjusted forage, pre-
pregnancy BMI, smoking,
smoking habits of partner,
drinking habits, gravidity,
parity, the number of cesarean
sections, uterine infection,
household income, educational
levels, and sex of infant
OR (95% CI)
Early preterm
Q1
Q2
Q3
Q4
Reference
0.66 (0.37, 1.20)
0.80 (0.46, 1.41)
1.22 (0.74, 2.02)
p for trend: 0.134
Late preterm
75th: 7.44 ng/g
Q1: Reference
Q2: 0.99 (0.78, 1.26)
Quartiles (ng/g)
Q3: 0.98 (0.77, 1.25)
Q1: <4.49
Q4: 0.92 (0.72, 1.18)
Q2: 4.80-5.95
p for trend: 0.920
Q3
Q4
5.96-7.43
>7.44
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RefereDCesignnd Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Goto etal. (2021)
Japan
January 2011 to
March 2014
Cohort
JECS
n: 15,540
First, data from participants
who withdrew from the study
or did not meet the Pb
measurement quality control
criteria were excluded
(n = 2,002). Second, data
from mothers who: were lost
to follow-up; had severe
maternal conditions
preceding pregnancy, such
as chronic hypertension,
pregestational diabetes or
cardiac disease, during
pregnancy; or had
pregnancies ending in
abortions or stillbirths
(n = 1,209) was excluded.
Third, data from infants with
chromosomal or major
congenital anomalies
(n = 263) or multiple births
(n = 283) was excluded.
Blood
Maternal blood measured
by ICP-MS
Age at measurement:
Maternal age at second or
third trimester (mean age at
delivery: 31 ± 5.0 yr)
Mean: 0.69 |jg/dL
Median: 0.63 |jg/dL
75th: 0.78 pg/dL
Max: 7.4 pg/dL
Preterm birth (<37
gestational weeks) risk
Preterm birth was defined
as a GA of less than 37
completed wk.
Age at outcome:
birth
Multivariable linear regression
models were adjusted for
maternal age at birth, BMI
before pregnancy, weight gain
during pregnancy, maternal
educational background, a
history of preterm birth, alcohol
consumption during pregnancy,
smoking habit during
pregnancy, and parity
OR (95% CI), per
0.1 pg/dL increase in
maternal blood Pb: 0.90
(0.70, 1.16)
OR (95% CI), per
doubling increment in
maternal blood Pb: 0.978
(0.689, 1.39)
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RefereDCesignnd Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Rabito etal. (2014)
Shelby County,
Tennessee
United States
2008-2011
Cohort
CANDLE study
n: 98
Healthy pregnant woman
between the ages of 16 and
40 yr, carrying a single fetus
with the intent to deliver the
fetus, residence within
Shelby County, Tennessee,
and having the intent to
deliver at one of three area-
based hospitals
Blood and cord blood
Maternal blood, collected at
second and third trimester
and at delivery, and cord
blood, collected at deliver,
were measured by ICP-MS
Age at measurement:
Maternal age at collection
(median: 29.50 yr)
Median:
Second trimester:
0.43 |jg/dL
Third trimester: 0.43 |jg/dL
At delivery: 0.50 |jg/dL
Cord blood: 0.37 |jg/dL
Preterm birth
Preterm birth (<37 wk),
early term birth (37-
39 wk), or full-term birth
(>39 wk) based on GA,
which was determined by
expected due data and
LMP
Age at outcome:
birth
Logistic regression models
were adjusted for marital
status, maternal education
level, and maternal income
Geometric mean (SD):
Second trimester: 0.42
(0.20) |jg/dL
Third trimester: 0.45
(0.28) |jg/dL
At delivery: 0.50
(0.35) |jg/dL
Cord blood: 0.37
(0.32) |jg/dL
Max:
Second trimester:
1.22 |jg/dL
Third trimester: 2.10 |jg/dL
At delivery: 2.47 |jg/dL
Cord blood: 1.80 |jg/dL
OR (95% Cl)b, per 0.1-
unit increase in maternal
blood Pb
Preterm birth
Second trimester: 1.66
(1.23, 2.23)
Third trimester: 1.24
(1.01, 1.52)
Early term birth
Second trimester: 0.87
(0.63, 1.20)
Third trimester: 0.88
(0.69, 1.13)
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RefereDCesignnd Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Taylor et al. (2015)
Bristol
UK
April 1991-December
1992
Cohort
ALSPAC
n: 4,285
All pregnant women in the
former Avon Health
Authority with an expected
delivery date between April
1, 1991, and December 31,
1992, were eligible for the
study
Blood
Maternal blood measured
by ICP-MS, collected as
early as possible in
pregnancy (median GA of
sampling: 11 wk)
Age at measurement:
Maternal age at
measurement
Preterm delivery
Preterm delivery was less
than 37 wk of gestation
Age at outcome:
birth
Logistic regression models for
preterm birth were adjusted for
maternal height, maternal pre-
pregnancy weight, maternal
educational attainment, parity,
number of cigarettes per day,
sex of baby
OR (95% Cl)b:
(1.35, 3.00)
2.00
Mean (SD): 3.67
(1.47) |jg/dL
Geometric mean:
3.43 |jg/dL
Median: 3.42 |jg/dL
Max: 19.14 pg/dL
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Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Li etal. (2017a)
China
January 1 to
December 31, 2009
Cohort
C-ABCS
n: 3,125
Mother-and-singleton-
offspring pairs from Hefei
City who provided informed
consent, did not drink
alcohol or smoke cigarettes
during pregnancy, did not
have mental disorders, did
not have pregnancy-induced
hypertension, preeclampsia,
gestational diabetes, heart
disease, thyroid-related
disease, a history of >3
previous miscarriages, or
plans to leave location
before delivery
Blood
Maternal blood (serum)
measured by GFAAS
coupled with a deuterium-
lamp background correction
system, collected in the first
and second trimesters
(median time for serum
collection: 14 gestational
weeks; range from 4 to 27
gestational week)
Mean: 1.50 |jg/dL
Max: 5.46 |jg/dL
Tertiles:
low-Pb: <1.18 |jg/dL
medium-Pb: 1.18-
1.70 |jg/dL
high-Pb: >1.71 pg/dL
Preterm birth
Gestational week was
calculated using mother's
last menstrual period.
Preterm birth was defined
as a live birth at less than
37 completed gestational and parity
weeks and preterm birth
can be further sub-divided
into early preterm birth
(<32 gestational weeks),
moderate preterm birth
(32 to <34 gestational
weeks) and late preterm
birth, 34 to <37
gestational weeks)
Age at outcome:
birth
Multiple logistic regression
models estimated the
association between maternal
serum Pb level and risk of
preterm birth, adjusted for
maternal age, pre-pregnancy
OR (95% CI):
Low-Pb: Reference
Medium-Pb: 2.33 (1.49,
3.65)
High-Pb: 3.09 (2.01,
BMI, monthly income, gravidity, 4.76)
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Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Ashrap et al. (2020)
Puerto Rico
2010-2017
Cohort
PROTECT
n: 731
Participants were recruited
at approximately 14 ± 2 wk
of gestation at seven
prenatal clinics and
hospitals throughout
Northern Puerto Rico and
followed until birth; maternal
age between 18 and 40 yr;
residence inside of the
Northern Karst aquifer
region; disuse of oral
contraceptives within the
3 mo prior to pregnancy;
disuse of IVF to become
pregnant; and free of any
major medical or obstetrical
complications, including pre-
existing diabetes. Each
woman participated in a total
of up to three study visits
(18 ± 2 wk, 22 ± 2 wk, and
26 ± 2 wk of gestation)
Blood
Maternal blood was
measured by ICP-MS
Age at measurement:
18-40 (collection between
18 and 26 wk of gestation)
Geometric mean (SD):
Preterm births: 0.39
(1.6) pg/dL
Term births: 0.32
(1.5) pg/dL
Median:
Preterm births: 0.36 pg/dL
Term births: 0.32 pg/dL
Preterm birth (overall and
spontaneous preterm
birth)
All the birth outcome data
were extracted from
medical records. Preterm
birth was defined as <37
completed weeks of
gestation with further
classification of
spontaneous preterm birth
(presentation of
premature rupture of the
membranes, spontaneous
preterm labor, or both)
and non-spontaneous
preterm birth (preterm
births with preeclampsia
or with both artificial
membrane rupture and
induced labor)
Age at outcome:
birth
Logistic regression models
were adjusted for maternal
age, maternal education level,
pre-pregnancy BMI, and
exposure to second-hand
smoking
OR (95% Cl)b, per IQR
increase in maternal
blood In-Pb
Preterm birth
Overall: 1.63 (1.17, 2.28)
Spontaneous: 1.53 (1.00,
2.35)
Tertilese:
Overall preterm birth:
T1: Reference
T2: 1.27 (0.65, 2.47)
T3: 1.93 (1.02, 3.62)
Spontaneous preterm
birth:
T1
T2
T3
Reference
0.69 (0.29, 1.66)
1.50 (0.71, 3.18)
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Design
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Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Ashrapet al. (2021) PROTECT
n = 682
Puerto Rico
2011-2017
Cohort
Participants were recruited
at approximately 14 ± 2 wk
of gestation at seven
prenatal clinics and
hospitals throughout
Northern Puerto Rico and
followed until birth; maternal
age between 18 and 40 yr;
residence inside of the
Northern Karst aquifer
region; disuse of oral
contraceptives within the
3 mo prior to pregnancy;
disuse of IVF to become
pregnant; and free of any
major medical or obstetrical
complications, including pre-
existing diabetes. Each
woman participated in a total
of up to three study visits
(18 ± 2 wk, 22 ± 2 wk, and
26 ± 2 wk of gestation)
Blood
Maternal blood was
measured by ICP-MS
Age at measurement:
18-40 (collection between
18 and 26 wk of gestation)
Geometric mean: 3.1 |jg/dL
Median: 3.1 |jg/dL
75th: 4.1 |jg/dL
95th: 6.5 pg/dL
Max: 15.1 pg/dL
Preterm birth (overall and Logistic regression models
spontaneous preterm
birth)
Birth outcomes were
extracted from medical
records. Psychosocial
status was evaluated
using four questionnaires
Age at outcome: birth
were adjusted for maternal
age, maternal education, pre-
pregnancy BMI, and exposure
to secondhand smoking
OR (95% Cl)b, per IQR
increase in in maternal
blood In-Pb
Preterm birth:
Good Psychosocial
Status: 1.72 (1.14, 2.58)
Poor Psychosocial
Status: 1.43 (0.69, 2.97)
Spontaneous preterm
birth:
Good Psychosocial
Status: 1.56 (0.93, 2.60)
Poor Psychosocial
Status: 1.22 (0.42, 3.56)
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Bui et al. (2022)
North Carolina
United States
2004-2009
Quasi-experimental
CCG MSA
n: 147,673 live births in the
CCG MSA; Treatment group
n: 1,138; Control group n:
13,398
Exogenous variation in Pb
exposure resulting from
NASCAR's deleading of
racing fuel in 2007 was used
as a quasi-experiment.
CMS, located in the CCG,
was the only NASCAR
racetrack in North Carolina
that held races every year
during our sample period.
Races occurred bi-annually,
in October and May,
ensuring that all full and
near full-term births in the
sample were prenatally
exposed via the mother to at
least one NASCAR event.
Births where the mother's
residential address was
within 4,000 meters of CMS
were classified as the
treatment group, while the
control group consists of
births where the mother's
residential address is in the
CCG but is at least
10,000m from the racetrack
centroid
Preterm birth
Preterm birth was defined
as clinical GA <37 wk.
Age at outcome: birth
Difference-in-difference models
were used to compare birth
outcomes in a non-randomized
treatment group before and
after NASCAR deleading to
those in the control group.
Models were adjusted for
mother's age, education, race,
and smoking behavior; father's
age, education, and race;
infant's birth order and sex; as
well as proximity to a TRI
facility or airport, median
household income, and age of
housing stock; a set of census
tract, month, and year indicator
variables were also included
(3 (95% Cl)b, as
estimated average
treatment effect of
treatment group
All births
Any exposure: -0.03
(-0.057, -0.002)
Trimester 1: -0.247
(-0.438, -0.057)
Trimester 2: 0.019
(-0.042, 0.079)
Trimester 3: -0.163
(-0.277, -0.049)
ALSPAC = Avon Longitudinal Study of Parents and Children; As = arsenic; BMI = body mass index; C-ABCS = China-Anhui Birth Cohort Study; CANDLE = Conditions Affecting Neurocognitive
Development and Learning in Early Childhood; CCG = Charlotte-Concord-Gastonia; Cd = cadmium; CMS = Charlotte Motor Speedway; Cr = chromium; EMASAR = Study on the Environment
and Reproductive Health; GA = gestational age; GFAAS = graphite furnace atomic absorption spectrometry; hr = hour(s); Hg = mercury; ICP-MS = inductively coupled plasma mass
spectrometry; INMA = Instituto de Nanociencia y Materiales de Aragon; IVF = in vitro fertilization; LBW = low birth weight; LMP = last menstrual period or last missed period; LOD = limit of
detection; mo = month(s); MSA = Metropolitan Statistical Area; OR = odds ratio; PROM = premature rupture of membranes; PROTECT = Puerto Rico Test site for Exploring Contamination
Threats; SD = standard deviation; TRI = Toxics Release Inventory; UCB = umbilical cord blood; wk = week(s); yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect estimates are
standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval. Categorical effect
estimates are not standardized.
bEffects estimates unable to be standardized.
°Pb measurements were converted from |jg/L to |jg/dL.
dPb measurements were converted from ng/mLto |jg/dL.
eNo cut points provided for the categorizations.
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Table 8-6
Epidemiologic studies of Pb exposure and birth defects
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Jin et al. (2013)
Shanxi Province
China
October 2002 -
onward
Case-control
n: 210: 80 controls, 50 any
NTD case; 36 cases of
anencephaly; and 44 cases
of spina bifida
Once a fetus with an NTD
was identified as a case, a
healthy newborn without
congenital malformations
was selected as a control.
The control was of the same
sex as the case and had a
mother residing in the same
county as that of the case. In
this study, we randomly
selected 36 cases of
newborns with anencephaly
and 44 cases of newborns
with spina bifida as case
groups and 50 healthy term
newborns as a control group.
Other: Placenta
Placental tissue, collected
at delivery or pregnancy
termination, was
measured with ICP-MS
Age at Measurement:
delivery or pregnancy
termination
Mean (SD)
Controls: 22.38 (16.35)
ng/g; NTD cases: 23.30
(22.42) ng/g;
Anencephaly cases:
19.30 (15) ng/g
Spina bifida cases: 23.04
(20.03) ng/g
Median
Controls: 16.9 ng/g
NTD cases: 17.59 ng/g
Anencephaly cases:
10.96 ng/g Spina bifida
cases: 17.38 ng/g
75th:
Controls: 28.83 ng/g
NTD cases: 28.15 ng/g
Anencephaly cases:
28.86 ng/g
Spina bifida cases:
28.86 ng/g
Birth defects: NTDs
Trained local health
workers made primary
diagnoses by physical
examination of the
fetal/newborn body for
any pregnancy
outcomes and filled in
a reporting form for
each case. Three
pediatricians
independently
reviewed the case
report forms and
photographs before
assigning the final
diagnostic codes
Age at outcome:
birth or pregnancy
termination
No attempt was made
to adjust for
confounding factors in
our analyses of Pb
because no differences
in their placental
concentrations were
present between cases
and controls.
OR (95% Cl)b:
Any NTD: 1.14 (0.56, 2.30)
Anencephaly: 1.08 (0.46,
2.56)
Spina bifida: 1.19 (0.53, 2.67)
Liu et al. (2021)
n: 332
Other: Umbilical cord
tissue
Birth defects: NTDs
Multivariate logistic
regression model
OR (95% Cl)b:
1.94)
1.23 (0.78,
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Shanxi, China
2004-2016
Case-control
Fetuses from elective
pregnancy terminations and
newborns from the Shanxi
Province in China. Cases
were defined as those with
NTD, and controls were
healthy newborns matched
by maternal residence and
date of last menstruation.
Umbilical cord tissue
measured by ICP-MS
Age at measurement:
At delivery or elective
termination
Median: 26.18 ng/g
75th: 48.58 ng/g
Max: 225.572 ng/g
NTD cases were
diagnosed by fetal
ultrasound scan or
physical examination
at birth or pregnancy
termination.
Age at outcome:
birth or pregnancy
termination
adjusted for folic acid
supplementation
Categorization:
Low exposure
(<1.10 ng/g)
High exposure
(>=1.10 ng/g)
Tian et al. (2021) n: 750
Shanxi province,
China
2003-2016
Case control
Participants were recruited
from six counties or cities in
the Shanxi province of
northern China.
Blood
Maternal blood (serum)
was measured by ICP-MS
Age at measurement:
Maternal age at collection
Median0:
Controls: 0.087 |jg/dL
Case: 0.115 |jg/dL
75thc:
Controls: 0.197 |jg/dL
Cases: 0.268 |jg/dL
Birth defects: NTDs
Diagnoses of
malformation are
made by local health
workers through
physical examination
of the newborns or
electively terminated
fetuses, in
combination with fetal
ultrasound scans.
Age at outcome: birth
or pregnancy
termination
Multilevel mixed effects
logistic regression
model adjusted for
maternal age, maternal
BMI, education,
gestational weeks, sex
of the fetus,
periconceptional folic
acid use, maternal flu,
or fever.
OR (95% CI):
Tertilesd
NTDs
Lowest: Reference
Medium: 2.05 (1.05, 4.02)
Highest: 3.51 (1.76, 6.98)
p for trend: <0.001
Spina bifida
Lowest: Reference
Medium: 2.16 (1.00, 4.88)
Highest: 5.16 (2.24,
p for trend: 0.022
Anencephaly
Lowest: Reference
11.87)
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Reference and
Study Design
Study Population Exposure Assessment Outcome
Confounders Effect Esti™*fs and 95%
Pi etal. (2018)
Shanxi Province
(Pingding, Xiyang,
Taigu, and Zezhou)
China
2005-2007
Case-control
Medium: 2.97 (1.09, 8.12)
Highest: 5.54 (1.89, 16.19)
p for trend: 0.002
Female Infants
NTDs
Lowest: Reference
Medium: 2.63 (0.99, 7.24)
Highest: 6.45 (2.20, 18.95)
p for trend: 0.001
Male Infants
NTDs
Lowest: Reference
Medium: 2.11 (1.02, 4.34)
Highest: 2.16 (1.03, 4.59)
p for trend: 0.048
n: 103 cases and 206
controls
Newborns or terminated
fetuses with any major
external structural defects,
including OFCs, NTDs,
congenital hydrocephalus,
limb defects were recruited
from five rural counties in
Shanxi Province
Other: Placenta
Placental tissue, collected
immediately after delivery,
was measured by ICP-MS
Age at Measurement:
birth
Mean (SD)
Controls 72.6 (34.8) ng/g
Case: 130.9 (95.7) ng/g
Median
Controls: 67.9 ng/g
Cases: 96.1 ng/g
75th
Controls: 98.1 ng/g
Cases: 176.4 ng/g
Birth defects: OFCs
Diagnoses of
newborns/fetuses with
major birth defects
were done through
physical examination
or prenatal ultrasound
examination by county
healthcare workers.
Once a newborn/fetus
with a major birth
defect was identified
as a case, a healthy
newborn with no
congenital
malformation was
selected as a control
to match the case by
residence of the
mother (the same
Binary logistic
regression adjusted for
occupation, newborn
sex, gestational weeks,
previous history of birth
defects, maternal flu or
fever, and passive
smoking during the
periconceptional period
OR (95% CI)
Orofacial defects:
T1: Reference
T2: 3.88 (1.78, 8.42)
T3: 5.17 (2.37, 11.29)
p for trend: <0.001
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Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Tertiles (ng/g):
T1: <57.5
T2: 57.5-96.8
T3: >96.8
county), date of the
LMP (±4 wk), and
newborn sex.
Age at outcome:
at diagnosis
Takeuchi et al.
(2022)
Japan
2011-2014
Case-control
JECS
n:192 cases, 1920 matched
controls
Pregnant women living in the
study area and
understanding of the
Japanese language.
Participants were excluded if
they had missing data
(heavy metal data, matching
variables, and/or both).
Covariates for matching
were maternal age,
psychological stress
measured by the K6 score,
gestational weeks of blood
sampling during second
trimester, folic acid intake
estimated from a food-
frequency questionnaire,
alcohol intake (self-reported),
smoking (self-reported),
education level, BMI before
pregnancy, diabetes before
pregnancy, intake of
supplements (self-reported),
and regional center
Blood
Maternal blood measured
by ICP-MS
Age at measurement:
Maternal age at collection
(second trimester)
Mediane
Cohort: 0.585 |jg/dL
Cases: 0.584 |jg/dL
Controls: 0.575 |jg/dL
75the
Cohort: 0.73 |jg/dL
Cases: 0.72 |jg/dL
Controls: 0.71 |jg/dL
Birth defects: Cleft
palate and cleft lip
(isolated)
Validated medical
records were used to
identify isolated cleft
lip and palate.
Conditional logistic
regression adjusted for
sex and concentrations
of Hg, Cd and Mn
OR (95% CI), per 0.1 pg/dL
increase in maternal blood
Pb: 1.10 (0.55, 2.21)
Mivashita et al.
(2021)
JECS.
N: 89,273
Blood
Birth defects:
Abdominal congenital
malformations
Multivariate logistic OR (95% CI)
regression models were Abdominal congenital
adjusted for maternal malformations
age, smoking habit,
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Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Japan
January 2011-
2014
Cohort
¦March
Pregnant women and their
newborns recruited for the
JECS. Singleton, live births
were included.
Maternal blood (serum)
measured by ICP-MS.
Age at measurement:
maternal age at collection
(mid-late pregnancy)
Median
Cohort: 5.84 ng/g
Controls: 5.85 ng/g
Cases: 5.53
75th
Cohort: 7.32 ng/g
Controls: 7.32 ng/g
Cases: 7.00 ng/g
Max
Cohort: 110 ng/g
Quartiles (ng/g):
Q1
Q2
Q3
Q4
<4.7
4.7-<5.84
5.84-<7.32
>7.32
Abdominal congenital
malformations
(including
omphalocele,
gastroschisis,
esophageal atresia
with/without fistula,
duodenal atresia,
intestinal atresia,
anorectal atresia,
diaphragmic hernia)
were identified from
birth records or
records 1 mo post
birth
Age at outcome: birth
to month post birth
drinking habit, paternal
smoking habit, birth
year of child, sex of
child
Q1
Q2
Q3
Q4
Reference
1.19 (0.76, 1.84)
0.77 (0.47, 1.26)
0.85 (0.52, 1.38)
p for trend: 0.233
Diaphragmic hernia
Q1
Q2
Q3
Q4
Reference
1.24 (0.51, 2.99)
0.89 (0.34, 2.31)
0.81 (0.30, 2.20)
p for trend: 0.543
Omphalocele
Q1
Q2
Q3
Q4
Reference
0.72 (0.29, 1.81)
0.35 (0.11, 1.12)
0.35 (0.11, 1.13)
p for trend: 0.033
Gastroschisis
Q1: Reference
Q2
Q3
Q4
1
1.00 (0.14, 7.09)
2.63 (0.50, 13.70)
p for trend: 0.212
Esophageal atresia
with/without fistula
Q1
Q2
Q3
Q4
Reference
0.49 (0.04, 5.43)
0.95 (0.13, 6.80)
1.88 (0.33, 10.50)
p for trend: 0.346
Duodenal atresia/stenosis
Q1
Q2
Q3
Reference
0.25 (0.03, 2.27)
0.50 (0.09, 2.75)
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Reference and
Study Design
Study Population Exposure Assessment Outcome
Confounders Effect Esti™*fs and 95%
Liu etal. (2018)
China
February 2010-
October2011
Case-control
Q4: 0.99 (0.24, 4.06)
p for trend: 0.910
Intestinal atresia/stenosis
Q1: Reference
Q2: 1.40 (0.31, 6.29)
Q3: 1.06 (0.21, 5.27)
Q4: 1.12 (0.22, 5.64)
p for trend: 0.989
Anorectal atresia/stenosis
Q1: Reference
Q2: 1.65 (0.74, 3.67)
Q3: 0.57 (0.19, 1.68)
Q4: 0.62 (0.21, 1.83)
p for trend: 0.158
n: 97 cases with CHDs and
201 controls without any
abnormalities
Cord blood
UCB (serum) was
measured by ICP-MS
Age at Measurement:
birth
Median0
Cases: 0.791 |jg/dL
Controls: 0.740 ug/dL
75thc
Case: 0.922 |jg/dL
Controls: 0.877 |jg/dL
Tertiles (|jg/dL):
Low: <0.696
Medium: 0.696-0.826
High: >0.826
Birth defects: CHDs
Cardiac defects
diagnosed during
prenatal examination
were recruited as the
case group.
Age at outcome: age
at diagnosis
Logistic regression
models were adjusted
for maternal age,
maternal pre-pregnancy
BMI, maternal
education level, folic
acid supplement, and
parental smoking
OR (95% CI)
CHD, Overall
Low: Reference
Medium: 1.46 (0.77, 2.77)
High: 1.67 (0.88, 3.17)
Septal Defects
Low: Reference
Medium: 1.20 (0.57, 2.52)
High: 1.61 (0.78, 3.32)
Conotruncal Defects
Low: Reference
Medium: 1.35 (0.60, 3.06)
High: 1.47 (0.65, 3.34)
Right-sided Outflow Tract
Deformity
Low: Reference
Eligible fetuses with cardiac
defects diagnosed during
prenatal examination were
recruited as the case group.
For each case, one pregnant
control without any fetal
malformation was selected in
the same hospital with a
gestation age within 2 wk of
the case fetus. Cases and
controls with GAs from 14 to
40 wk were selected for this
study after the following
exclusion criteria were
applied: (1) multiple
pregnancies; (2) CHD family
history; (3) fetus diagnosed
with a chromosomal
abnormality or hereditary
syndrome; (4) fetus with
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Reference and
Study Design
Study Population Exposure Assessment Outcome
Confounders Effect Esti™*fs and 95%
extra cardiac malformations;
(5) uncompleted
questionnaire for some
reason. CHD cases were
classified into six subtypes
based on the anatomic
lesion: (i) septal defects, (ii)
conotruncal defects, (iii) left
sided outflow tract deformity,
(iv) right-sided outflow tract
deformity, (v) anomalous
pulmonary venous return,
and (vi) other cardiac
structural abnormalities.
Medium: 0.92 (0.37, 2.26)
High: 1.21 (0.50, 2.94)
Left-sided Outflow Tract
Deformity
Low: Reference
Medium: 2.29 (0.62, 8.41)
High: 1.32 (0.29, 5.91)
Anomalous Pulmonary
Venous Return
Low: Reference
Medium: 1.71 (0.37, 7.83)
High: 1.49 (0.30, 7.44)
Other Cardiac Structural
Abnormalities
Low: Reference
Medium: 1.10 (0.36, 3.40)
High: 1.41 (0.47, 4.22)
BMI = body mass index; Cd = cadmium; CHD = congenital heart diseases/defects; GA = gestational age; Hg = mercury; ICP-MS = inductively coupled plasma mass spectrometry;
K6 = Kessler Psychological Distress Scale; LMP = last menstrual period or last missed period; Mn = manganese; mo = month(s); NTD = neural tube defect; OFC = orofacial cleft;
OR = odds ratio; SD = standard deviation; UCB = umbilical cord blood; wk = week(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bEffects estimates unable to be standardized.
°Pb measurements were converted from ng/mL to |jg/dL.
dNo cut points provided for the categorizations.
ePb measurements were converted from |jg/L to |jg/dL.
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Table 8-7 Epidemiologic studies of Pb exposure and fetal and infant mortality and spontaneous abortion and
pregnancy loss
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
Xu et al. (2012)
Guiyu and Xiamen
China
2001-2008
Cross-sectional
n: 531 (n = 432 from Guiyu
and n = 99 from Xiamen)
Women who gave birth in
Guiyu or non-urban area of
Xiamen between 2001 and
2008
Cord blood
UCB measured by
GFAAS
Age at Measurement:
birth
Median:
Guiyu: 10.78 pg/dL
Xiamen: 2.25 |jg/dL
Max:
Guiyu: 47.46 |jg/dL
Xiamen: 7.22 |jg/dL
Stillbirth rate
Stillbirth was defined as
fetal death before complete
expulsion or extraction from
the mother at >20 wk of
gestation
Age at outcome:
birth
Multiple logistic regression
models were adjusted for
maternal age and infant sex
OR (95% Cl)b: 4.20
(3.40, 5.18)
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Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
Louis etal. (2017)
Michigan and Texas
United States
2005-2009
Cohort
LIFE Study
n: 344
Female partners aged 18-40
and male partners aged
>18 yr who were in a
committed relationship; no
physician diagnosis of
infertility/sterility; off
contraception <2 mo; and an
ability to communicate in
English or Spanish. Female
partners also had to have
menstrual cycles ranging
between 21 and 42 d as
required by the fertility monitor
and without the use of
injectable hormonal
contraceptives in the past year
given the uncertain timing for
ovulation return
Blood
Blood from female and
male partners was
measured by ICP-MS
Age at Measurement:
18-40 for females and
>18 for males
Median
Females: 0.66 |jg/dL
Males: 1.00 |jg/dL
75th:
Females: 0.82 |jg/dL
Males: 1.37 |jg/dL
Pregnancy loss
Pregnancy was
prospectively captured by
women's use of the
Clearblue® digital home
pregnancy test, which is
sensitive in detecting
25 mlU/mL ofhCGand
accurately used by women.
Depending upon timing of
loss, it was detected by
conversion to a negative
pregnancy test, clinical
confirmation, or return of
menses.
Age at outcome:
18-40 yr
Cox proportional hazard
models; individual partner
model adjusted for age, BMI,
history of prior loss conditional
on gravidity, average number
of daily alcoholic drinks
consumed, and cigarettes
smoked during the
preconception and early
pregnancy windows for
females and preconception for
males; couples based model
adjusted for each partner's
metal concentration, age,
difference in couples' ages,
BMI, average number of daily
alcoholic drinks consumed and
cigarettes smoked during the
preconception and early
pregnancy window for females
and preconception for males,
and history of prior loss
conditional on gravidity
HR (95% Cl)b
Individual partner
model
Female partner:
1.01 (0.82, 1.25)
Male partner: 0.95
(0.77, 1.17)
Couple based
model
Female partner:
1.01 (0.80, 1.28)
Male partner: 0.96
(0.77, 1.22)
Viaeh et al. (2021)
Tehran
Iran
March 2016-October
2017
Cohort
Tehran Environment and
Neurodevelopmental Disorder
n: 166 (spontaneous abortion
n: 25 and ongoing pregnancy
n: 141)
Pregnant women with GA of
10-16 wk and of Iranian
nationality and Tehran city
inhabitant were invited to
participate in the study.
Blood
Maternal blood was
measured using ICP-MS
Age at measurement:
maternal age at first
trimester
Mean0: 4.96 |jg/dL
Maxc: 70.982 |jg/dL
Spontaneous abortion
Spontaneous abortion
defined as fetal demise
before 20 wk gestation and
reported by study
participant or research
hospital.
Age at outcome: before
20 wk of gestation
Logistic regression models
adjusted for maternal age,
primipara, and previous
abortion
OR (95% CI), per
0.1 |jg/dL increase
in maternal blood
Pb: 1.08 (0.98, 1.20)
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Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates
and 95% Clsa
Tolunav et al. (2016) n: 101
Ankara
Turkey
January 2012 and July
2012
Cohort
The study group consisted of
patients with ongoing
pregnancy (n = 20) and the
reference group consisted of
patients experienced ART
failure, miscarriage, or
biochemical pregnancy
(n = 81)
Blood
Maternal blood was
measured by AAS
Age at Measurement:
20-40
Median
Study group: 2.34 |jg/dL
Reference: group
5.11 |jg/dL
Max
Study group: 7.97 |jg/dL
Reference group:
10.47 |jg/dL for
reference group
Pregnancy loss
Clinical pregnancy was
defined as the presence of
an embryo with a heartbeat
at 6th gestational week.
Ongoing pregnancy was
defined when the
pregnancy had completed
20 wk of gestation.
Implantation rate was
calculated separately for
each woman as the number
of gestational sacs divided
by the number of
transferred embryos
multiplied by 100.
Age at outcome:
completion of 20 wk of
gestation
Log binominal regression
analysis adjusted for age and
BMI
RR (95% CI): 0.978
(0.957, 0.999)
Li et al. (2022)
Hefei
China
October 2019 -
January 2020
Cohort
n: 1184
Participants were selected
from First Affiliated Hospital of
Anhui Medical University while
seeking IVF treatment and
diagnosed with infertility with
their partner. Inclusion criteria:
women were aged between
20 and 45 yr; couples were
diagnosed with infertility
(failure to establish a clinical
pregnancy with unprotected
intercourse for at least 1 yr);
and IVF indicators were tubal
factor, ovulation failure, or
other factors for female
Blood
Maternal blood (serum)
was measured by ICP-
MS
Age at measurement:
maternal age at
collection (day before
oocytes were retrieved
for IVF); mean age was
30.22 yr
Geometric meand:
0.0877 |jg/dL
Mediand: 0.0924 |jg/dL
Spontaneous abortion
Spontaneous abortion
before gestational week 12
was followed upon the 65th
day after embryo transfer.
Age at outcome: maternal
age at outcome (before
gestational week 12)
Logistic regression model for
successful implantation
adjusted for: maternal age,
BMI, treatment protocol,
numbers of retrieved oocytes,
embryo quality
OR (95%CI)b:
Spontaneous
abortion: 1.39 (1.02,
1.91)
Tertiles
Low: Reference
Medium: 1.49 (0.84,
2.63)
High: 1.55 (0.87,
2.79)
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Outcome
Confounders
Effect Estimates
and 95% Clsa
partner or male factor or
unexplained fertility.
75thd: 0.14399 pg/dL
Tertilesd (pg/dL):
Low: 0.002-0.065
Medium: 0.065-0.125
High: 0.125-0.481
AAS = atomic absorption spectrometry; BMI = body mass index; d = day(s); GFAAS = graphite furnace atomic absorption spectrometry; hCG = human chorionic gonadotropin;
HR = hazard ratio; ICP-MS = inductively coupled plasma mass spectrometry; IVF = in vitro fertilization; mo = month(s); OR = odds ratio; UCB = umbilical cord blood; wk = week(s);
yr = year(s).
aEffect estimates are standardized to a 1 pg/dL increase in blood Pb or a 10 pg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect estimates are
standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval. Categorical
effect estimates are not standardized.
bEffects estimates unable to be standardized.
°Pb measurements were converted from pg/L to pg/dL.
dPb measurements were converted from ng/Lto pg/dL.
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Table 8-8
Epidemiologic studies of Pb exposure and placental function
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Al-Saleh et al. (2014) n: 1,578
Al-Kharj
Saudi Arabia
2005-2006
Cross-sectional
Women aged 16-50 yr
who delivered in Al-
Kharj hospital, Saudi
Arabia
Blood, cord blood, and
other: placenta
Maternal blood, UCB, and
placental tissue
measured by AAS
Age at Measurement:
maternal age 16-50; birth
Mean ± SD:
Maternal blood:
2.897 ± 1.851 |jg/dL
UBC:
2.551 ±2.592 pg/dL
Placenta:
0.579 ±2.176 pg/g
Median:
Maternal blood:
2.540 pg/dL
UCB: 2.057 pg/dL
Placenta: 0.450 pg/g
75th:
Maternal blood:
3.314 pg/dL
UCB: 2.689 pg/dL
Placenta: 0.630 pg/g
Max:
Maternal blood:
25.955 pg/dL
UCB: 56.511 pg/dL
Placenta: 78 pg/g
Placental function:
Placental thickness
Placental weight and
placental thickness were
recorded by obstetrician in
delivery room
Age at outcome:
birth
Logistic regression model was
adjusted for maternal age, parity,
mother's third trimester BMI,
urinary cotinine, mother's highest
education, total family income,
and GA
OR (95% Cl)b, per unit
increase in maternal blood
Pb: 1.64 (1.12, 2.41)
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Confounders
Effect Estimates and 95%
Clsa
Tsuii etal. (2019)
Japan
January 2011-March
2014
Cross-sectional
JECS
n: 16,019
Mothers who delivered
a singleton pregnancy
Blood
Maternal blood, collected
during the second
trimester, was measured
by ICP-MS
Age at Measurement:
maternal age at second
trimester
Median: 5.96 ng/g
75th: 7.45 ng/g
Quartiles:
Q1: <4.79 ng/g
Q2: 4.80-5.95 ng/g
Q3: 5.96-7.44 ng/g
Q4: >7.45 ng/g
Placental function:
Placenta previa and
placenta accreta
Data for those with and
without placenta previa
and placenta accreta were
obtained from medical
records.
Age at outcome:
maternal age at diagnosis
Multivariable logistic regression
models were adjusted for age,
smoking, smoking habits of the
partner, drinking habits, gravidity,
parity, number of cesarean
deliveries, and geographic region;
Placenta previa was added as a
covariate when comparisons
were performed with or without
placenta accreta
OR (95% CI):
Placenta previa
Q1
Q2
Q3
Q4
Reference
2.59 (1.40, 4.80)
1.32 (0.66, 2.64)
1.34 (0.67, 2.67)
p for trend: 0.007
Placenta accreta
Q1
Q2
Q3
Q4
Reference
1.46 (0.57, 3.76)
1.68 (0.66, 4.24)
0.79 (0.27, 2.30)
p for trend: 0.345
AAS = atomic absorption spectrometry; BMI = body mass index; CI = confidence interval; GA = gestational age; ICP-MS = inductively coupled plasma mass spectrometry; JECS = Japan
Environment and Children's Study; OR = odds ratio; Q = quartile; SD = standard deviation; UCB = umbilical cord blood; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect estimates are
standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval. Categorical effect
estimates are not standardized.
bEffect estimates unable to be standardized.
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Table 8-9 Epidemiologic studies of Pb exposure and other pregnancy and other birth outcomes
RefereDCesignnd Study Population Exposure Assessment Outcome Confounders EffeCt9E5%1c$*
Ashley-Martin et al.
(2015a)
Vancouver, Edmonton,
Winnipeg, Sudbury,
Ottawa, Kingston,
Toronto, Hamilton,
Montreal, and Halifax
Canada
2008-2011
Cohort
MIREC study
n: 1,260
Women were recruited from
10 Canadian sites during
their first trimester and
consented to provide urine
and blood samples. Women
were eligible for inclusion if
they were <14 wk gestation
at the time of recruitment,
>18 yr of age, able to
communicate in French or
English, and planning to
deliver at a local hospital
Blood
Maternal blood was
measured by ICP-MS
Age at Measurement:
Maternal age during 1st
and 3rd trimester
Geometric mean (SD):
0.88 (1.61) [jg/dL
Quartiles (pg/dL):
Q1: <0.63
Q2
Q3
Q4
0.64 to <0.87
0.88 to <1.20
>1.20
Other Pregnancy and Birth
Outcomes: Fetal metabolic
function
Leptin and adiponectin were
measured in plasma from
1363 stored UCB samples by
ELISA using kits from Meso
Scale Discovery. All samples
were above the LOD.
Age at outcome:
birth
Logistic regression models
were adjusted for maternal
age at delivery, pre-
pregnancy BMI, parity, and
BWZ
OR (95% CI)
Low leptin and maternal
blood Pb:
Q1
Q2
Q3
Q4
Reference
0.9 (0.5, 1.6)
0.6 (0.3, 1.1)
0.9 (0.5, 1.5)
High leptin and maternal
blood Pb:
Q1
Q2
Q3
Q4
Reference
1.2 (0.7, 2.1)
1.0 (0.6, 1.8)
1.7 (1.0, 2.9)
Low adiponectin and
maternal blood Pb:
Q1
Q2
Q3
Q4
Reference
1.3 (0.8, 2.2)
0.8 (0.5, 1.4)
1.1 (0.6, 1.9)
High adiponectin and
maternal blood Pb:
Q1: Reference:
Q2: 0.9 (0.5, 1.5)
Q3: 1.1 (0.7, 1.9)
Q4: 0.9 (0.5, 1.5)
Herlin et al. (2019) n: 194 enrolled of the 221 Blood, cord blood, and Other Pregnancy and Birth Multivariable-adjusted (3 (95% Cl)c:
pregnant women other: placenta Outcomes: rTL linear regression models; UCB'-0 038 (-0 074
Salta Province (Andean models with maternal blood _g qo2)
All pregnant women living in The rTL was measured in Pb were adjusted for
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Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
part)
Argentina
October 2012-
December 2013
Cohort
the Andean part of the Salta
province northern Argentina
with estimated delivery date
between October 2012 and
December 2013, were
invited to participate
Maternal blood, UCB,
and placenta were
measured using ICP-MS
Age at Measurement:
birth
Median:
Maternal bloodb:
2.1 |jg/dL
UCBb: 1.4 Mg/dL
Placenta: 5.8 |jg/kg
Max:
Maternal bloodb:
9.9 |jg/dL
UCBb: 6.0 |jg/dL
Placenta: 38 |jg/kg
maternal blood leukocytes
(blood samples collected in
late pregnancy, mainly third
trimester), cord blood
leukocytes, and placental
tissue. We obtained high-
quality DNA and measured
rTL in 169 blood samples of
the pregnant women, 99 of
their placentas, and 98 cord
blood samples of their
babies. The rTL was
measured as the ratio
between the signal intensity
of the telomere sequences
and the signal intensity of a
single-copy gene
(hemoglobin (3 chain), using
real-time polymerase chain
reaction.
Age at outcome:
birth
maternal age, pre-
pregnancy BMI, and
education; models with
placenta were also
adjusted for GA at birth;
models with UCB were
adjusted for maternal age,
pre-pregnancy BMI, GA at
birth, and BW.
Maternal blood: 0.026
(-0.043, 0.095)
Placenta: -0.029 (-0.074,
0.016)
Liao etal. (2015)
Taiwan
n: 113
Pregnant women were
recruited from a single
March-December 2010 institution in northern
Taiwan
Cross-sectional
Blood
Maternal blood
(plasma), collected at
the first trimester
(between 10 and 14wk
of gestation), was
measured by ICP-MS
Age at Measurement:
Maternal age at first
trimester (mean age
30.92 ±3.09 yr)
Geometric mean:
0.048 |jg/L
Other Pregnancy and Birth
Outcomes: Fetal nuchal
translucency thickness
Fetal nuchal translucency
thickness was measured at
10-14 wk of gestation by a
gynecologist and three
trained sonographers
Age at outcome:
Age at scan (between
gestational week 10 and 14)
Multiple linear regression
models were adjusted for
maternal age, gestational
weeks, pre-pregnancy BMI,
supplement use, and
medication
(3 (95% Cl)c: 0.022 mm
(-0.06, 0.10)
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Outcome
Confounders
Effect Estimates and
95% Clsa
Ashley-Martin et al.
(2015b)
Canada
2008-2011
Cohort
MIREC Study
n: 1256
Pregnant women in Canada
who had singleton, term
birth (>37 wk)
Blood
Maternal blood was
measured by ICP-MS
Age at measurement:
maternal age at first and
third trimester
Median: 0.62 |jg/dL
75th: 1.03 pg/dL
Max: 4.14 pg/dL
Other Pregnancy and Birth
Outcomes: Fetal immune
system biomarkers
Immune system biomarkers
were measured in the
plasma of UCB samples;
TSLP concentrations were
determined using a
commercial antibody kit; IL-
33 concentrations were
analyzed using antibodies
from an R & D systems duo
set; IgE was determined from
ELISA kits
Logistic regression
adjusted for maternal age
OR (95% CI)
Maternal log-io-Pb blood
concentrations with elevated
(>80%) cord blood
concentrations of IL-33 and
TSLP: 0.79 (0.62, 1.01)
Maternal log-io-Pb blood
concentrations with elevated
(>0.5 kU/L) cord blood
concentrations of IgE: 0.99
(0.77, 1.26)
Age at outcome:
birth
Taylor et al. (2014)
Bristol
UK
April 1991-December
1992
Cohort
ALSPAC study
n: 4,285
Pregnant women enrolled in
the ALSPAC study at a
median GA of 11 wk
Blood
Maternal blood was
measured by ICP-MS
Age at Measurement:
Maternal age at
measurement (median
GA of sampling: 11 wk)
Median:
Quintile 1
Quintile 2
Quintile 3
Quintile 4
Quintile 5
Max:
2.11 pg/dL
2.82 pg/dL
3.43 pg/dL
4.13 pg/dL
5.00 pg/dL
Other Pregnancy and Birth
Outcomes: Secondary sex
ratio
The sex of the infant was
recorded at birth
Age at outcome:
birth
Logistic regression models
adjusted for maternal and
paternal age, and parity
OR (95% CI)
Q1
Q2
Q3
Q4
Q5
Reference
1.04 (0.86, 1.42)
0.90 (0.70, 1.15)
1.01 (0.79, 1.30)
1.06 (0.82, 1.37)
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Effect Estimates and
95% Clsa
Q1: 2.53 |jg/dL
Q2: 3.11 |jg/dL
Q3: 3.71 |jg/dL
Q4: 4.63 pg/dL
Q5: 19.14 |jg/dL
Bloom etal. (2015)
Michigan (4 counties)
and Texas (12 counties)
United States
2005-2009
Cohort
LIFE
n: 235
Potential participants were
identified, using fishing
license registries or a
commercially available
direct marketing data base,
from 12 counties in Texas
and four in Michigan,
respectively, with presumed
exposure to persistent
organic pollutants. Inclusion
criteria comprised a
committed heterosexual
relationship, women aged
18-40 yr (men >18), English
or Spanish speaker, no use
of an injectable
contraceptive within 12 mo,
and a menstrual cycle
length of 21^2 d.
Blood
Maternal and paternal
blood, collected before
pregnancy (baseline),
were measured by ICP-
MS
Age at Measurement:
>18, maternal mean
age: 29.75 (SD: 3.73) yr
and paternal mean age:
31.52 (SD:4.57) yr
Mean (SD):
Maternal: 0.71
(0.30) pg/dL
Paternal: 1.13
(0.63) pg/dL
Median:
Maternal: 0.66 pg/dL
Paternal: 0.98 pg/dL
Max:
Maternal: 2.23 pg/dL
Paternal: 6.43 pg/dL
Other Pregnancy and Birth
Outcomes: Secondary sex
ratio
Women were followed until
delivery when they
completed and returned birth
announcements that
captured date and sex of
birth, weight and length, and
HC. Secondary sex ratio is
the ratio of live male to
female births, reflecting a
male excess.
Age at outcome:
birth
Log-binomial models for
secondary sex ratio: effect
of maternal exposure
adjusted for paternal
exposure, maternal age,
difference in maternal and
paternal age, and maternal
and paternal smoking,
income, race, serum lipids
(mg/dL), and creatinine for
urine (mg/dL); effect of
paternal exposure adjusted
for maternal exposure,
paternal age, difference in
maternal and paternal age,
and maternal and paternal
smoking, income, race,
serum lipids (mg/dL), and
creatinine for urine (mg/dL)
RR (95% CI)
Maternal Exposure:
T1: Reference
T2: 0.97 (0.78, 1.22)
T3: 1.00 (0.81, 1.24)
p for trend: 0.884
Paternal Exposure:
T1
T2
T3
Reference
1.12 (0.89, 1.41)
1.06 (0.84, 1.34)
p for trend: 0.854
Tertiles (pg/dL):
Maternal Blood Pb
T1: <0.55 (<33rd
percentile)
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Outcome
Confounders
Effect Estimates and
95% Clsa
T2: 0.55-0.73 (33rd to
67th percentile)
T3: >0.73 (>67th
percentile)
Paternal Blood Pb
T1: <0.84 (<33rd
percentile)
T2: 0.84-1.16 (33rd to
67th percentile)
T3: >1.16 (>67th
percentile)
Tatsuta et al. (2022b)
Japan
January 2011-March
2014 (followed through
birth)
Cohort
JECS
n: 85,171
Pregnant women and their
paternal partners were
recruited from 15 regions of
Japan. Participants
delivered a live infant with
singleton pregnancy and
had child sex information.
Participants were excluded
is they had a stillbirth,
abortion, multiple births, or
withdrew before birth;
missing blood sample
information; missing
confounders; or without
partner's consent and with
paternal age or occupational
exposure to Pb deficits
Blood
Maternal blood was
measured by ICP-MS.
Age at measurement:
maternal age at
collection (middle or late
pregnancy)
Median: 5.85 ng/g
Max: 110 ng/g
Quartiles (ng/g)
Q1
Q2
Q3
Q4
1.20-4.46
4.47-5.39
5.40-6.35
6.36-7.76
Other Pregnancy and Birth
Outcomes: Secondary sex
ratio
Sex of the infant obtained
from the medical record
transcripts by physicians,
midwives, nurses, or trained
research coordinators.
Logistic regression models
were adjusted for maternal
age at parturition, season
of birth, pre-pregnancy
BMI, annual household
income, gravidity, fertility
treatments, score of the
K6, maternal smoking
status during pregnancy,
passive smoking status
during pregnancy, birth
year and study area
(regional center)
OR (95% CI)
Q1
Q2
Q3
Q4
Q5
Reference
1.082 (1.037, 1.129)
1.122 (1.074, 1.171)
1.214 (1.163, 1.268)
1.279 (1.224, 1.336)
Q5: 7.77-110
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Effect Estimates and
95% Clsa
AAS = atomic absorption spectrometry; ALSPAC = Avon Longitudinal Study of Parents and Children; BMI = body mass index; BW = birth weight; BWZ = birth weight Z-score; d = day(s);
ELISA = enzyme-linked immunosorbent assay; HC = head circumference; ICP-MS = inductively coupled plasma mass spectrometry; IgE = immunoglobulin E; IL-33 = interleukin-33;
JECS = Japan Environment and Children's Study; K6 = Kessler Psychological Distress Scale; LIFE = Longitudinal Investigation of Fertility and the Environment; LOD = limit of detection;
MIREC = Maternal-Infant Research on Environmental Chemicals; mo = month(s); OR = odds ratio; Q = quartile; RR = relative risk; rTL = relative telomere length; SD = standard deviation;
T# = fertile #; TSLP = thymic stromal lymphopoietin; UCB = umbilical cord blood; wk = week(s); yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect estimates are
standardized to the specified unit increase for the 10,h-90,h percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval. Categorical effect
estimates are not standardized.
bPb measurements were converted from |jg/L to |jg/dL.
°Effect estimates unable to be standardized.
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Table 8-10
Epidemiologic studies of Pb exposure and postnatal growth
Reference and
Study Design
Study Population Exposure Assessment Outcome Confounders
Effect Estimates and 95%
Cisa
Sianes-Pastor et al.
(2021)
United States
2013-2016
Cross-sectional
NHANES
n: 1,634
Children aged 6-11 yr
old participating in the
2013-2014 and 2015-
2016 NHANES cycles
Blood
Blood was measured by
ICP-MS
Age at measurement: 6-
11 yrold
Median:
Overall: 0.5 |jg/dL
Girls: 0.5 |jg/dL
Boys: 0.5 |jg/dL
75th:
Overall: 0.8 |jg/dL
Girls: 0.7 |jg/dL
Boys: 0.8 |jg/dL
Max:
Overall: 5.8 |jg/dL
Girls: 5.8 |jg/dL
Boys: 5.0 |jg/dL
Postnatal growth: weight,
WC, upper arm length,
standing height, and BMI
Physical examination was
performed to obtain body
measurements.
Age at outcome: 6-11 yr
old
Linear regression
models were adjusted
for total calorie intake,
race, PIR, children's
age, smoker(s) in the
household, outside-of-
school and at-school
activity scores,
children's sex, and co-
exposure to fluoride,
Mn, Hg, and Se
(3 (95% CI)
BMI (kg/m2): -2.092 (-3.227,
-0.957)
Standing height (cm): -3.116
(-5.03, -1.202)
WC (cm): -5.742 (-8.769,
-2.715)
Upper arm length (cm): -1.068
(-1.625, -0.512)
Girls
BMI (kg/m2): -3.204 (-5.654,
-0.754)
Standing height (cm): -2.89
(-6.691, 0.911)
WC (cm): -6.659 (-12.911,
-0.408)
Upper arm length (cm): -1.696
(-2.859, -0.534)
Boys
BMI (kg/m2): -1.959 (-3.45,
-0.467)
Standing height (cm): -3.828
(-6.588, -1.068)
WC (cm): -6.81 (-10.995,
-2.626)
Upper arm length (cm): -0.89
(-1.691, -0.089)
Kuana et al. (2020) n: 395 Blood Postnatal growth: height, General linearized (3 (95% CI)
weight, bust, waistline, models were adjusted Height (cm)' -3 21 (-4 24
Nanjing Students aged 7-11 yr , ,, and BMI for age and gender -9-m
China (grades 2 to 4) were Blood was measured by
ICP-MS
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Effect Estimates and 95%
Cisa
2012
Cross-sectional
recruited from public
primary schools in
Nanjing, an industry city
from East China.
Students with congenital
mental retardation (third-
degree relatives
included) and other
serious diseases were
excluded. Students and
their parents were
informed of the research
content and purpose.
Only completely
matched groups of
samples, including
questionnaire
information, blood
samples, growth indexes
and school
performances, were
included in the study.
Age at Measurement:
7-11 yr
Mean (SD)b: 3.04
(1.72) pg/dL
Medianb: 2.61 pg/dL
Growth: Individual
measurements were
carried out by the medical
staff according to the
standard protocols of
WHO. Height was
measured using a
mechanical height gauge
to the nearest 0.1 cm.
Weight was measured
using digital scales to the
nearest 100 g.
Age at outcome:
7-11 yr
Weight (kg): -1.96 (-3.11,
-0.82)
Bust (cm): -2.77 (-3.79, -1.76)
Waistline (cm): -3.65 (-4.78,
-2.52)
BMI (kg/m2): -0.20 (-0.65,
0.25)
Zhou et al. (2020) n: 1,678
Taizhou Children 6 yr or older
China
April 2013-
November 2013
Cross-sectional
Blood
Blood was measured by
GFAAS
Age at Measurement:
>6 yr
Meanb: 5.684 pg/dL
Geometric meanb:
4.904 pg/dL
Medianb: 4.644 pg/dL
75thb: 6.4 pg/dL
Maxb: 46.8 pg/dL
Tertilesb (pg/dL)
Postnatal growth: HAZ,
WAZ and BMIZ
Children's body weight
and supine length or
standing height were
measured. BMI was
calculated by the formula
BMI = weight (kg)/height
(m)2; Z-scores of
anthropometric
parameters, such as HAZ,
WAZ and BMIZ, were
calculated with the WHO
Child Growth Standards.
Multivariable linear
models were adjusted
for age, sex, BW,
maternal education
(3 (95% Cl)c:
WAZ: -0.33 (-0.56, -0.11)
HAZ: -0.38 (-0.63, -0.14)
BMIZ: -0.13 (-0.37, 0.12)
WAZ Tertiles
T1
T2
T3
Reference
-0.28 (-0.47, -0.09)
-0.42 (-0.62, -0.23)
HAZ Tertiles
T1
T2
T3
Reference
-0.26 (-0.47, -0.04)
-0.36 (-0.58, -0.15)
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Effect Estimates and 95%
Cisa
T1: <2.5
T2: 2.5-5.0
T3: >5.0
Age at outcome:
>6 yr
BMIZ Tertiles
T1: Reference
T2
T3
-0.18 (-0.39, 0.04)
-0.29 (-0.50, -0.07)
Males:
WAZ: -0.36 (-0.67, -0.06)
HAZ: -0.38 (-0.72, -0.04)
BMIZ: -0.15 (-0.49, 0.19)
WAZ Tertiles:
T1
T2
T3
Reference
-0.42 (-0.71, -0.13)
-0.52 (-0.81, -0.24)
HAZ Tertiles:
T1
T2
T3
Reference
-0.36 (-0.69, -0.004)
-0.43 (-0.75, -0.11)
BMIZ Tertiles:
T1: Reference
T2: -0.28 (-0.60, 0.04)
T3: -0.35 (-0.68, -0.03)
Females
WAZ: -0.29 (-0.61, 0.03)
HAZ: -0.35 (-0.71, 0.01)
BMIZ: -0.10 (-0.45, 0.26)
WAZ Tertiles:
T1: Reference
T2: -0.17 (-0.42, 0.09)
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Cisa
T3
-0.36 (-0.62, -0.09)
HAZ Tertiles:
T1
Reference
T2
-0.17 (-0.45, 0.11)
T3
-0.31 (-0.60, -0.02)
BMIZ Tertiles:
T1
Reference
T2
-0.10 (-0.38, 0.18)
T3
-0.25 (-0.54, 0.04)
Choi etal. (2017) n: 210
Seoul
South Korea
July 2014 to June
2016
Cross-sectional
Children ranging from 8
to 23 mo in age and
healthy; no intake of
herbal medicine, iron, or
zinc supplements in the
past 3 mo; no acute
febrile disease or acute
gastrointestinal disease
in the past 2 wk; and no
evidence of other acute
or chronic diseases
affecting growth on
physical examination or
in medical history
Blood
Blood was measured by
ICP-MS
Age at Measurement:
8-23 mo
Geometric mean:
0.96 |jg/dL Median:
0.83 |jg/dL
75th: 1.23 |jg/dL
Max: 3.5 |jg/dL
Postnatal growth: Weight,
height, HC
Each infant's weight,
height, and HC were
measured by experienced
nurse; iron deficiency and
iron deficiency anemia,
complete blood count,
serum iron and ferritin
concentrations, as well as
total iron-binding capacity
were measured from the
venous blood samples of
infants
Age at outcome:
8-23 mo
Linear regression
models; BW,
sociodemographic and
feeding-related factors,
and iron and anemia
status
(3 (95% CI):
WAZ-BWZ (difference of the
WAZ at the time of the study
and BWZs): -0.238 (-0.391,
-0.085)
HCAZ: -0.213 (-0.366, -0.06)
Ashley-Martin et al.
(2019)
Vancouver,
Edmonton,
Winnipeg, Sudbury,
Ottawa, Kingston,
MIREC Study
n: 449
MIREC study is a
national-level pregnancy
cohort of 2001 women
from 10 cities across
Blood
Blood was measured by
ICP-MS
Age at Measurement:
2-5 yr
Postnatal growth: HAZ,
WAZ, BMIZ
Child anthropometry was
performed during the
home visit and served as
a measure of growth at
Linear regression
models adjusted for
maternal education,
maternal country of
birth, age, postnatal
BMI, maternal prenatal
smoking, and paternal
(3 (95% Cl)b
HAZ
Overall:
T1: Reference
T2: -0.015 (-0.23, 0.20)
8-168
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
Toronto, Hamilton,
Montreal, and
Halifax
Canada
2008-2011
Cross-sectional
Canada including
Vancouver, Edmonton,
Winnipeg, Sudbury,
Ottawa, Kingston,
Toronto, Hamilton,
Montreal, and Halifax.
Participants were
recruited in the first
trimester of pregnancy
between 2008 and 2011
and followed through
delivery.
Median: 0.663 [jg/dL
75th: 0.962 pg/dL
Max: 5.49 pg/dL
Tertiles (pg/dL)
T1
T2
T3
<0.54
0.54-0.82
>0.82
that time. Weight and
height were measured
using a calibrated scale
and calibrated
stadiometer. All
measurements were
completed in duplicate or,
if warranted due to
predefined differences in
duplicate measurements,
in triplicate.
Age at outcome:
2-5 yr
BMI; models were
additionally adjusted for
maternal metal
concentrations
T3: 0.025 (-0.20, 0.25)
Male
T1
T2
T3
Reference
0.003 (-0.28, 0.29)
-0.039 (-0.32, 0.24)
Female
T1
T2
T3
Reference
0.022 (-0.31, 0.35)
0.095 (-0.26, 0.45)
WAZ
Overall
T1
T2
T3
Reference
0.064 (-0.12, 0.25)
-0.004 (-0.20, 0.19)
Male
T1
T2
T3
Reference
0.11 (-0.15, 0.36)
0.074 (-0.18, 0.33)
Female
T1
T2
T3
Reference
0.050 (-0.22, 0.32)
-0.11 (-0.40, 0.18)
BMIZ
Overall
T1: Reference
T2: 0.097 (-0.098, 0.29)
T3: -0.041 (-0.24, 0.16)
Male
T1: Reference
T2: 0.15 (-0.13, 0.42)
8-169
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Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
T3: 0.14 (-0.14, 0.41)
Female
T1
T2
T3
Reference
0.039 (-0.24, 0.32)
-0.26 (-0.55, 0.033)
Adjusted for maternal
exposure:
HAZ
Overall:
T1
T2
T3
Reference
-0.030 (-0.25, 0.19)
-0.008 (-0.25, 0.23)
Male
T1
T2
T3
Reference
-0.007 (-0.30, 0.28)
-0.067 (-0.38, 0.24)
Female
T1
T2
T3
Reference
0.013 (-0.33, 0.36)
0.081 (-0.30, 0.46)
WAZ
Overall
T1
T2
T3
Reference
0.041 (-0.15, 0.23)
-0.05 (-0.26, 0.16)
Male
T1
T2
T3
Reference
0.09 (-0.16, 0.35)
0.04 (-0.24, 0.32)
8-170
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
Female
T1
T2
T3
Reference
0.024 (-0.26, 0.30)
-0.15 (-0.46, 0.17)
BMIZ
Overall
T1
T2
T3
Reference
0.076 (-0.12, 0.28)
-0.086 (-0.30, 0.14)
Male
T1
T2
T3
Reference
0.14 (-0.14, 0.41)
0.11 (-0.19, 0.41)
Female
T1
T2
T3
Reference
0.006 (-0.28, 0.29)
-0.32 (-0.64, 0.0036)
Jedrvchowski et al.
(2015)
Krakow
Poland
January 2001-
February 2004
Cohort
Krakow Cohort Study
n: 379
The present analysis
was restricted to 379
term-babies (born
>36 wk of gestation)
who took part in the 9-yr
follow-up. Women who
were residents of
Krakow, one of the
major cities in Poland,
and attended
ambulatory prenatal
clinics in the first and
second trimesters of
pregnancy were eligible
Blood and cord blood
Maternal and UCB,
obtained at delivery, and
blood (capillary),
obtained at age 5, were
measured by high-
performance liquid
chromatography
atmospheric-pressure
ionization tandem mass
spectrometry
Age at Measurement:
Maternal age at delivery
and 5 yr old
Postnatal growth: Height
gain
At ages of 3-9 children
were invited annually for
pediatric examination
during which height
measurements were
done.
Age at outcome:
3-9 yr old
GEE models were
adjusted for maternal
height, BL, pre-
pregnancy maternal
weight, gestational
weight gain, prenatal
and postnatal ETS,
breastfeeding, maternal
education, and parity
(3 (95% CI), as mean height
growth (cm) by UCB tertiles
T1
T2
T3
Reference
-0.671 (-1.610, 0.267)
-0.736 (-1.779, 0.307)
8-171
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Reference and
Study Design
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Outcome
Confounders
Effect Estimates and 95%
Cisa
for the study. Enrollment
included only
nonsmoking women with
singleton pregnancies
between the ages of 18
and 35 yr who were free
from such chronic
diseases as diabetes
and hypertension.
Geometric mean:
UCB: 1.21 |jg/dL
Blood: 2.05 pg/dL
UBC Tertiles (pg/dL)
T1
T2
T3
<1.0
1.1-1.4
>1.4
Kimetal. (2017)
Korea
January 2011-
December 2012
Cohort
CHECK
n: 280
Healthy pregnant
women with mature term
singleton were recruited,
who did not have
preterm delivery,
medical predisposition,
or history of
occupational exposure
Cord blood
UCB was measured by
GFAAS
Age at measurement:
birth
Mean:
Overall: 1.31 pg/dL
Males: 1.39 pg/dL
Females: 1.21 pg/dL
Postnatal growth: Weight, Generalized linear
height, and BMI
Weight and height were
measured by the health
professionals
Age at outcome:
3, 6, 9, 12, 15, 18, 24, and
27 mo of age
model adjusted for
maternal age, maternal
BMI, gestational period,
cesarean section, and
smoking
(3 (95% Cl)b
Weight
At birth: 0.037 (-0.128, 2.01)
3 mo: -0.039 (-0.414, 0.335)
6 mo: -0.391 (-0.814, 0.033)
9 mo: 0.000 (-0.356, 0.357)
12 mo: 0.125 (-0.302, 0.552)
15 mo: 0.093 (-0.396, 0.582)
18 mo: 0.897 (-0.171, 1.965)
24 mo: 0.717 (0.195, 1.239)
27 mo: 0.316 (-0.345, 0.977)
Height
At birth: 0.176 (-0.003, 0.354)
3 mo: -0.023 (-0.384, 0.337)
6 mo: 0.033 (-0.458, 0.523)
9 mo: 0.049 (-0.346, 0.444)
12 mo: -0.058 (-0.531, 0.415)
15 mo: 0.226 (-0.220, 0.671)
18 mo: 0.909 (-0.222, 2.040)
24 mo: 0.138 (-0.530, 0.806)
27 mo: 0.354 (-0.497, 1.205)
8-172
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
BMI
At birth: -0.167 (-0.357, 0.023)
3 mo: -0.019 (-0.431, 0.392)
6 mo: -0.461 (-0.937, 0.014)
9 mo: -0.031 (-0.430, 0.369)
12 mo: -0.020 (-0.492, 0.452)
15 mo: -0.098 (-0.481, 0.285)
18 mo: 0.157 (-1.266, 1.580)
24 mo: 0.695 (0.077, 1.313)
27 mo: 0.409 (-0.398, 1.216)
Males
Weight
At birth: 0.088 (-0.140, 0.316)
3 mo: -0.008 (-0.597, 0.581)
6 mo: -0.023 (-0.543, 0.497)
9 mo: 0.167 (-0.398, 0.733)
12 mo: 0.202 (-0.631, 1.034)
15 mo: 0.365 (-0.467, 1.197)
18 mo: 1.324 (0.023, 2.626)
24 mo: 0.962 (0.181, 1.743)
27 mo: 0.417 (-0.631, 1.465)
Height
At birth: 0.270 (0.037, 0.502)
3 mo: 0.232 (-0.262, 0.726)
6 mo: -0.077 (-0.695, 0.540)
9 mo: 0.166 (-0.363, 0.695)
12 mo: -0.147 (-1.153, 0.859)
15 mo: 0.433 (-0.147, 1.013)
18 mo: 1.648 (0.270, 3.026)
24 mo: 1.062 (-0.132, 2.255)
8-173
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Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
27 mo: 1.618 (-0.450, 3.686)
BMI
At birth: -0.194 (-0.413, 0.025)
3 mo: -0.130 (-0.800, 0.540)
6 mo: 0.003 (-0.558, 0.563)
9 mo: -0.009 (-0.522, 0.504)
12 mo: 0.314 (-0.689, 1.318)
15 mo: -0.049 (-0.569, 0.470)
18 mo: 0.319 (-1.496, 2.135)
24 mo: 0.472 (-0.172, 1.116)
27 mo: 0.966 (-1.390, 3.322)
Females
Weight
At birth: 0.006 (-0.236, 0.248)
3 mo: -0.072 (-0.640, 0.496)
6 mo: -0.828 (-1.502, -0.154)
9 mo: -0.098 (-0.602, 0.407)
12 mo: 0.101 (-0.443, 0.644)
15 mo: -0.039 (-0.722, 0.643)
18 mo: -0.826 (-15.627,
13.976)
24 mo: 0.821 (-0.087, 1.728)
27 mo: 0.236 (-1.089, 1.561)
Height
At birth: 0.102 (-0.177, 0.381)
3 mo: -0.249 (-0.875, 0.378)
6 mo: 0.106 (-0.732, 0.945)
9 mo: 0.104 (-0.526, 0.734)
12 mo: -0.057 (-0.608, 0.493)
15 mo: 0.121 (-0.664, 0.905)
18 mo: -0.788d
8-174
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
24 mo: -0.176 (-1.225, 0.874)
27 mo: -0.153 (-1.405, 1.100)
BMI
At birth: -0.142 (-0.474, 0.189)
3 mo: 0.098 (-0.491, 0.687)
6 mo: -0.974 (-1.778, -0.170)
9 mo: -0.143 (-0.805, 0.519)
12 mo: -0.147 (-0.688, 0.393)
15 mo: -0.103 (-0.712, 0.505)
18 mo: -2.263d
24 mo: 1.108 (-0.147, 2.362)
27 mo: 0.439 (-1.581, 2.460)
Hong etal. (2014)
Seoul, Cheonan,
and Ulsan
South Korea
May 2006 to
December 2010
Cohort
MOCEH
n: 1,751
This research was
conducted as a part of
MOCEH, which is a
multicenter prospective
hospital and community-
based birth cohort study.
Women who lived in
these cities were
enrolled in the first
trimester. The
participants fulfilled the
inclusion criterion of age
>18 yr. Written informed
consent was obtained at
the initial visit from all
enrolled mothers on
behalf of themselves
and their children. The
study subjects were
restricted to those in
which maternal and cord
BLLs were assessed,
Blood and cord blood
Maternal blood, obtained
during early pregnancy
(before gestational week
20) and at delivery, and
UCB were measured by
AAS
Age at Measurement:
maternal age at week 20
and at delivery; delivery
Mean:
Early pregnancy:
1.25 |jg/dL Late
pregnancy: 1.25 |jg/dL
UCB: 0.91 |jg/dL
Median:
Early pregnancy:
1.29 |jg/dL
Postnatal growth: weight
Z-score, length z-cores
Weights and lengths at 6
and 12 mo were taken by
using an infantometer by
laying infants on the
center of a scale and were
read to 1 decimal place
for weight (0.1 kg) and
length (0.1 cm). At 24 mo
of age, weights and
lengths were obtained by
using an automatic
measuring station for
weight and length by
standing
on the center of the scale
on both feet, and placing
their heels, bottom, back,
and posterior head on the
measuring rod.
Multivariable regression
models were adjusted
for mother's age,
education, pre-
pregnancy BMI, GA,
gender of the child, and
clinic location, and
calcium intake
(3 (95% Cl)b
Maternal Blood: Early
pregnancy Pb
Weight Z-scores
At birth: -0.05 (-0.16, 0.07)
6 mo: -0.03 (-0.19, 0.13)
12 mo: -0.10 (-0.26, 0.06)
24 mo: -0.05 (-0.23, 0.12)
Length Z-scores
At birth: 0.01 (-0.15, 0.18)
6 mo: -0.17 (-0.37, 0.02)
12 mo: 0.04 (-0.15, 0.24)
24 mo: -0.15 (-0.35, 0.04)
Maternal Blood: Late
Pregnancy Pb
Weight Z-scores
At birth: -0.01 (-0.15, 0.12)
6 mo: -0.15 (-0.34, 0.03)
12 mo: -0.15 (-0.34, 0.03)
8-175
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Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
and postnatal growth
measurements were
performed. Exclusion
criteria: LBW(<2500 g);
preterm birth
(gestational week <37);
and missing information
on maternal age, BMI,
education level, and
gestational week;
subjects with >2 SD for
mean maternal BLLs
and child BW or length
Late pregnancy:
1.27 |jg/dL
UCB: 0.93 pg/dL
75th:
Early pregnancy:
1.65 pg/dL
Late pregnancy:
1.64 pg/dL
UCB: 1.19 pg/dL
Max:
Early pregnancy:
2.63 pg/dL
Late pregnancy:
2.52 pg/dL
UCB: 1.90 pg/dL
Age at outcome:
6, 12 and 24 mo
24 mo: -0.33 (-0.53, -0.13)
Length Z-scores
At birth: -0.07 (-0.25, 0.11)
6 mo: -0.05 (-0.28, 0.16)
12 mo: 0.10 (-0.12, 0.33)
24 mo: -0.30 (-0.53, -0.08)
UCB Pb
Weight Z-scores
At birth: 0.08 (-0.04, 0.21)
6 mo: 0.10 (-0.07, 0.28)
12 mo: 0.06 (-0.10, 0.24)
24 mo: -0.01 (-0.21, 0.18)
Length Z-scores
At birth: 0.14 (-0.03, 0.32)
6 mo: 0.11 (-0.11, 0.33)
12 mo: 0.22 (0.01, 0.44)
24 mo: 0.004 (-0.22, 0.22)
Renzetti et al.
(2017)
Mexico City
Mexico
July 2007-February
2011
Cohort
PROGRESS
n: 513
Women were
considered eligible for
enrollment if they were
18 yr or older, pregnant
at <20 wk of gestation,
free of heart or kidney
disease, did not use
steroids or anti-epilepsy
drugs, did not consume
alcohol on a daily basis,
had access to a
telephone, and planned
to reside in Mexico City
for the following 3 yr
Blood, cord blood, and
bone
Maternal blood, collected
in the second and third
trimester of pregnancy
and within 12 hr of
delivery, and UCB,
collected within 12 hr of
delivery, were measured
by ICP-QQQ.
Maternal bone,
measured at 1-mo
postpartum from tibia
(cortical bone) and
patella (trabecular bone),
Postnatal growth: HAZ,
WAZ, BMIZ, and
percentage body fat
Trained research
assistants collected
measures of
anthropometry at the age
4-6-yr visit in which child
weight and standing
height were measured
using a professional
digital scale. BMI was
calculated from height and
weight and to determine
BMIZ for age and sex
based on WHO norms.
Multivariable linear
regression adjusted for
mother's age, BMI
(height when the
outcome is HAZ),
education, GA (weeks),
primiparity, smoke
exposure, delivery
mode, breastfeeding,
sex of the child, food
frequency questionnaire
total dietary intake,
LeadCare childhood
blood Pb, and child's
age (when the outcome
is percent body fat)
(3 (95% Cl)c
HAZ
Maternal blood, second
trimester: -0.04 (-0.13, 0.04)
Maternal blood, third trimester:
-0.10 (-0.19, -0.01)
Maternal blood, at delivery:
-0.04 (-0.13, 0.05)
UCB: -0.04 (-0.14, 0.06)
Maternal patella: 0.01 (-0.003,
0.02)
Maternal tibia: -0.003 (-0.01,
0.01)
WAZ
8-176
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
were measured using a
K-XRF instrument
Age at Measurement:
Maternal age at second
and third trimester and at
birth; child's age at
follow-up (4-6 yr)
Mean (SD):
Maternal blood - second
trimester: 3.7 (2.6) [jg/dL
Maternal blood - third
trimester: 3.9 (2.8) [jg/dL
Maternal blood - at
delivery: 4.3 (3.1) [jg/dL
UCB: 3.5 (2.7) pg/dL
Patella: 4.7 (8.8) pg/g
Tibia: 2.9 (8.6) pg/g
Geometric mean:
Maternal blood - second
trimester: 3.0 pg/dL
Maternal blood - third
trimester: 3.1 pg/dL
Maternal blood - at
delivery: 3.5 pg/dL
UCB: 2.8 pg/dL
Max:
Maternal blood - second
trimester: 17.8 pg/dL
Maternal blood - third
trimester: 28.3 pg/dL
Maternal blood - at
delivery: 21.9 pg/dL
Tetrapolar bioelectrical
impedance was measured
to estimate body fat mass
and percent body fat
Age at outcome:
4-6 yr old
Maternal blood, second
trimester: -0.02 (-0.13, 0.09)
Maternal blood, third trimester:
-0.11 (-0.22, -0.003)
Maternal blood, at delivery:
-0.03 (-0.13, 0.08)
UCB: -0.03 (-0.15, 0.09)
Maternal patella: 0.01 (-0.01,
0.02)
Maternal tibia: -0.0003 (-0.01,
0.01)
BMIZ
Maternal blood, second
trimester: 0.04 (-0.07, 0.15)
Maternal blood, third trimester:
-0.01 (-0.12, 0.10)
Maternal blood, at delivery:
-0.03 (-0.08, 0.14)
UCB: 0.05 (-0.08, 0.17)
Maternal patella: 0.01 (0.01,
0.02)
Maternal tibia: 0.01 (-0.01,
0.02)
Percentage of body fat
Maternal blood, second
trimester: -0.13 (-0.75, 0.49)
Maternal blood, third trimester:
-0.21 (-0.82, 0.41)
Maternal blood, at delivery:
-0.12 (-0.74, 0.50)
UCB: 0.31 (-0.37, 0.99)
Maternal patella: 0.01 (-0.06,
0.07)
8-177
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
UCB: 18.5 |jg/dL
Patella: 43.2 |jg/g
Tibia: 30.1 |jg/g
Maternal tibia: 0.01 (-0.06,
0.08)
Liu etal. (2019a)
Mexico City
Mexico
1994-2003
Cohort
ELEMENT
n: 248
Pregnant women who
were recruited from
three maternity hospitals
in Mexico City and
followed for 12 mo post-
partum and children
followed through age 4
Blood and bone
Maternal tibia (cortical)
and patella (trabecular)
bone, measured at 1-mo
postpartum, were
measured using a
noninvasive spot-source
Cd K-XRF instrument
constructed at Harvard
University. Blood,
obtained from each child
annually from 1 to 4 yr,
was measured by
GFAAS
Age at Measurement:
maternal age at delivery
and 1-4 yr old
Mean:
Maternal patella:
12.3 |jg/g
Maternal tibia: 8.9 |jg/g
Blood (cumulative):
19.6 |jg/dL
Median:
Maternal patella:
10.6 |jg/g
Maternal tibia: 8.3 |jg/g
Blood (cumulative):
17.7 |jg/dL
75th:
Postnatal growth: BMIZ,
WC, sum of skinfolds, and
body fat percentage
At the follow-up visit, child
weight, height, WC, and
skinfold thickness (biceps,
subscapular and
suprailiac) were measured
Age at outcome:
8-16 yr old
Multivariable linear
regression models were
adjusted for maternal
age, parity, education
and calcium treatment
group, and children's
age, sex, and pubertal
stage
(3 (95% Cl)c
BMIZ
Patella: -0.02 (-0.03, -0.01)
Tibia: -0.00 (-0.02, 0.01)
Blood: 0.02 (-0.40, 0.45)
WC (cm)
Patella: -0.12 (-0.22, -0.03)
Tibia: -0.07 (-0.21, 0.07)
Blood: -0.38 (-3.74, 2.97)
Sum of skinfolds (mm)
Patella: -0.29 (-0.50, -0.08)
Tibia: -0.10 (-0.38, 0.19)
Blood: -1.62 (-8.76, 5.52)
Body of fat percentage (%)
Patella: -0.09 (-0.17, -0.01)
Tibia: -0.01 (-0.13, 0.10)
Blood: 2.08 (-0.98, 5.13)
8-178
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
Maternal patella:
19.7 |jg/g
Maternal tibia: 15.2 |jg/g
Blood (cumulative):
23.5 |jg/dL
Max:
Maternal patella:
50.1 |jg/g
Maternal tibia: 38.6 |jg/g
Blood (cumulative):
55.0 |jg/dL
Afeiche et al. (2012) n: 773
Mexico City
Mexico
1994-2005
Cohort
Mothers were recruited
from maternity hospitals
serving low-to-moderate
income populations in
Mexico City; preterm
(<37 wk) and LBW
(<2500 g) were
excluded
Blood and bone
Maternal bone, assessed
at approximately 1 mo
postpartum, measured
by in vivo K-XRF from
the mid-tibial shaft
(cortical bone) and the
patella (trabecular bone);
blood, obtained from
children at 24 mo or 30-
48 mo, was measured by
GFAAS.
Age at Measurement:
Maternal age 1 mo
postpartum, with average
age at delivery: 25.7 (SD:
5.3) yr; birth—24 mo; 30-
48 mo
Postnatal growth: Attained
height and BMI
Children's weight and
height were measured
and recorded by trained
staff members at birth and
age 48 mo using standard
protocols
Age at outcome:
birth and 48 mo
Linear regression
models were adjusted
for maternal height and
calf circumference,
number of previous
pregnancies, marital
status, education level,
breastfeeding for 6 mo,
cohort, calcium
treatment group
assignment during
lactation and
pregnancy, age at
delivery, and child sex
and GA at birth; all
height models were
additionally adjusted for
BL; BMI models were
additionally adjusted for
BW
(3 (95% CI)
Height differences (cm)
Prenatal: -4.6 (-10.25, 1.05)
Infant blood: -0.84 (-1.43,
-0.26)
Childhood blood: 0.41 (-0.17,
0.99)
BMI difference (kg/m2)
Prenatal: -0.70 (-3.05, 1.65)
Infant blood: -0.07 (-0.32,
0.18)
Childhood blood: 0.09 (-0.15,
0.33)
Median:
Maternal tibia: 8.2 |jg/g
Maternal patella: 9.4 |jg/g
Infant blood (average
8-179
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
from birth to 24 mo)
4.5 |jg/dL
Childhood blood
(average from 30-
48 mo): 5.6 |jg/dL
Kerretal. (2019) n: 538
Torreon
Mexico
February 2001-
June 2002
Cohort
Children attending nine
public elementary
schools located within a
3.5 km radius from a
foundry close to the city
center participated in the
study. Participants were
randomized into one of
four groups: iron (30 mg
of ferrous fumarate),
zinc (30 mg zinc oxide),
a combination of iron
and zinc or a placebo
(sugar pill)
Blood
Blood, collected at
baseline (T1), 6 mo after
baseline (T2), and 12 mo
after baseline (T3), was
measured by GFAAS
Age at Measurement:
6-8 yr old
Median: 10.1 pg/dL
75th: 23.7 pg/dL
Postnatal growth: height,
knee height, and HAZ
A single trained individual
took anthropometric
measures at each time
point (T1, T2, T3),
according to standard
methods recommended
by the WHO; Height and
knee height were
measured without shoes
using a standardized
measuring board or a
knemometer, respectively,
to the nearest 1 mm;
Age at outcome:
6-8 yr old
Multivariable linear
regression adjusted for
age, sex, mother's
education, crowding,
and hemoglobin at
baseline; HAZ models
were not adjusted for
age or sex; models
were also stratified by
ALAD genotype
(3 (95% CI)
Height: -0.11 cm ( -0.18,
-0.04)
Knee height: -0.04 cm (-0.07,
-0.02)
HAZ: -0.02 cm ( -0.03, -0.01)
ALADl-2/2-2
Height: -0.38 cm (-0.68, 0.09)
Knee height: -0.14 cm (-0.25,
-0.02)
HAZ: -0.07 (-0.12, -0.02)
ALAD1-1
Height: -0.09 cm (-0.16,
-0.02)
Knee height: -0.04 cm (-0.06,
-0.01)
HAZ: -0.02 (-0.03, -0.004)
Burns et al. (2017)
Chapaevsk
Russia
2003-2005 (2012—
2015)
Cohort
Russian Children's
Study
n: 499
The Russian Children's
Study is a prospective
cohort of 499 boys
residing in Chapaevsk,
Russia, enrolled in
2003-2005 at ages 8-
9 yr and followed
Blood
Blood measured by
GFAAS with Zeeman
background corrected
Age at Measurement:
8-9 at enrollment
Postnatal growth: HAZ
and BMIZ
At study entry and annual
follow-up visits, a
standardized
anthropometric
examination was
performed according to a
written protocol. Height
was measured to the
Mixed effects linear
regression models were
adjusted for BW,
preterm birth, percent
calories from protein at
baseline, and age for
the HAZ models and
BW, no biological father
in home, percent
calories from fat at
(3 (95% Cl)b, as estimated
mean growth Z-scores
comparing higher (>5 pg/dL) to
lower (<5 pg/dL) BLL
HAZ: -0.43 (-0.60, -0.25)
BMIZ: -0.22 (-0.45, 0.006)
8-180
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
annually through 2012-
2015 to age
18 yr. For this analysis,
10 boys in the original
cohort were excluded
due to chronic illnesses
that could affect growth
and/or pubertal
development.
Median: 3.0 |jg/dL
Max: 31 |jg/dL
nearest 0.1 cm using a
stadiometer. Weight was
measured to the nearest
100 g with a metric scale.
HAZ and BMIZ were
calculated using the WHO
standards
Age at outcome:
8-9 at enrollment and
annually through age 18
baseline, and age for
BMIZ models
Deierlein et al.
(2019)
New York City, NY;
Cincinnati, OH; and
San Francisco, CA
United States
2004-2007
Cohort
Breast Cancer and
Environment Research
Program
n: 683
Girls ages 6-8 yr were
enrolled in 2004-2007
at three sites: New York
City, Cincinnati, and San
Francisco; girls have no
underlying endocrine
medical conditions, be
of Black or Hispanic
race/ethnicity (New York
City site only), and have
been born in the Kaiser
Permanente system
(san Francisco)
Blood
Blood was measured by
ICP-MS
Age at Measurement:
6-10 yr
Median: 0.99 |jg/dL
Mean (SD): 1.16
(0.67) |jg/dL
Geometric mean:
1.03 |jg/dL (95% CI:
0.99, 1.07)
Max: 5.40 |jg/dL
Postnatal growth: height,
BMI, WC, and percent
body fat
Weight (kg), standing
height (cm), and umbilical
WC (cm) were collected at
baseline and at biannual
(Cincinnati) or annual
(New York City and San
Francisco Bay Area)
follow-up visits by trained
interviewers using a
standard protocol; BMI
was calculated as weight
divided by squared height
(kg/m2). Percent body fat
was estimated using
bioelectrical impedance
analysis
Age at outcome:
7-14 yr
Linear mixed effects
models with an
unstructured correlation
matrix were adjusted for
age, age squared, race,
an Interaction term
between age and blood
Pb concentrations, an
interaction term
between age squared
and blood Pb
concentrations, and an
interaction term
between race and age
(3 (95% CI)
Height (cm)
Age
7:
-2.0 (-
-3.0,
-1.0)
Age
8:
-1.9 (-
-2.8,
-0.9)
Age
9:
-1.7 (-
-2.7,
-0.8)
Age
10
-1.6
(-2.6
-0.7)
Age
11
-1.6
(-2.5,
-0.6)
Age
12
-1.5
(-2.5,
-0.5)
Age
13
-1.5
(-2.5,
-0.5)
Age
14
-1.5
(-2.5,
-0.4)
BMI
(kg/m2)
Age
7:
-0.7 (-
-1.2,
-0.2)
Age
8:
-0.8 (-
-1.3,
-0.3)
Age
9:
-0.9 (-
-1.4,
-0.4)
Age
10
-0.9
(-1.4,
-0.4)
Age
11
-0.9
("1.5,
-0.3)
Age
12
-0.9
(-1.5,
-0.3)
Age
13
-0.8
(-1.5,
-0.2)
Age
14
-0.8
(-1.5,
-0.02)
WC (cm)
Age 7:
-2.2 (-
-3.8,
-0.6)
Age
8:
"2.5 (-
-3.8,
-1.1)
8-181
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
Age 9:
Age 10
Age 11
Age 12
Age 13
Age 14
Percent
Age 7
Age 8
Age 9
Age 10
Age 11
Age 12
Age 13
Age 14
¦2.7 (-4.0, -1.4)
-2.9 (-4.9, 1.4)
-3.0 (-4.5, -1.4)
-3.0 (-4.7, -1.3)
-3.0 (-4.8, -1.1)
-2.9 (-4.8, -0.9)
body fat (%)
¦1.8 (-3.2, -0.4)
¦2.0 (-3.3, -0.7)
¦2.1 (-3.4, -0.8)
-2.2 (-3.4, -0.9)
-2.1 (-3.4, -0.9)
-2.1 (-3.4, -0.8)
-1.9 (-3.2, -0.6)
-1.7 (-3.1, -0.4)
Raihan et al. (2018) MAL-ED study
n: 729
Mirpur, Dhaka
Bangladesh
Children under the age
of 2
November 2009-
December 2012
Cross-sectional
Blood
Blood was measured
using GFAAS
Age at measurement:
under the age of 2
Mean: 8.25 |jg/dL
Postnatal growth:
Stunting, wasting,
underweight
Child's length and weight
were measured using
Seca 417 infantometer
(precision: ± 1 mm) and
Seca 354 Dual Purpose
Baby Scale (precision:
10 gm).
Age at outcome:
under the age of 2
Logistic regression
models were adjusted
for child's gender,
weight, maternal
education, BMI, average
household income and
HFIAS categories in
stunting models; child's
gender, age, maternal
education, BMI, average
household income and
HFIAS categories in the
wasting models; and
child's gender, length,
maternal education,
BMI, average household
income and HFIAS
categories in the
underweight models
OR (95% CI)
Stunting: 1.78 (1.07, 2.99)
Wasting: 1.18 (0.64, 2.19)
Underweight: 1.63 (1.02, 2.61)
8-182
-------
Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Cisa
Gleason et al.
(2016)
Sirajdikhan and
Pabna Upazilas
Bangladesh
2008-2011 (2010-
2013)
Cohort
n: 618
Children of mother's
from Sirajdikhan and
Pabna Upazilas of
Bangladesh between
2008 and 2011;
Between 2010 and
2013, when children
were aged 12 to 40 mo,
healthcare workers from
Dhaka Community
Hospital invited families
to enroll their children in
follow-up studies
Cord blood
UCB were measured by
ICP-MS and child's
blood, collected at 20 to
40 mo, was measured by
portable LeadCare II
instruments
Age at Measurement:
at birth and 12-40 mo
Median:
UCB: 3.1 |jg/dL
Blood: 4.2 |jg/dL
75th:
UCB: 6.3 |jg/dL
Blood: 7.6 |jg/dL
Postnatal growth: Stunting
Stunting status of children
was determined using the
WHO macros (Version
3.2.2)
Age at outcome:
12-40 mo
Logistic regression
models were adjusted
for maternal weight,
maternal education,
maternal protein intake,
and HOME Inventory
score were all modeled
as continuous variables;
average water As and
Mn levels were included
as continuous variables
OR (95% CI)
UCB: 0.97 (0.93, 1.00)
Blood at 20-40 mo: 1.15 (1.00,
1.33)
ALAD = 6-aminolevulinic acid dehydratase; BL = birth length; BMI = body mass index; BMIZ = BMI-for-age Z-score; BW = birth weight; BWZ = birth weight Z-score;
CHECK = Children's Health and Environmental Chemicals in Korea; CI = confidence interval; ELEMENT = Early Life Exposure in Mexico to Environmental Toxicants;
ETS = environmental tobacco smoke; GEE = generalized estimating equation; GFAAS = graphite furnace atomic absorption spectrometry; HAZ = height-for-age Z-score;
HCAZ = head circumference for age Z-score; HFIAS = Household Food Insecurity Access Scale; HOME = Health Outcomes and Measures of the Environment; hr = hour(s); ICP-
MS = inductively coupled plasma mass spectrometry; ICP-QQQ = inductively coupled plasma triple quad; K-XRF = K-shell X-ray fluorescence; LBW = low birth weight; MAL-
ED = Interactions of Malnutrition and Enteric Infections: Consequences for Child Health and Development; MIREC = Maternal-Infant Research on Environmental Chemicals;
mo = month(s); MOCEH = Mothers' and Children's Environmental Health; NHANES = National Health and Nutrition Examination Survey; OR = odds ratio; PIR = poverty-income
ratio; PROGRESS = Programming Research in Obesity, Growth, Environment and Social Stressors; SD = standard deviation; T# = fertile #; UCB = umbilical cord blood;
WAZ = weight for age Z-score; WC = waist circumference; WHO = World Health Organization; wk = week(s); yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bPb measurements were converted from |jg/L to |jg/dL.
°Effect estimates unable to be standardized.
dNo CI reported.
8-183
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Table 8-11
Animal toxicological studies of Pb exposure and development
Study
Species (Stock/Strain), n, Sex
Timing of Exposure
Exposure
Details
(Concentration,
Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
Graham et al. (2011) Rat (Sprague-Dawley)
Control (vehicle), M/F, n = 14-16 (7-8/7-8)
1 mg/kg Pb, M/F, n = 14-16 (7-8/7-8)
10 mg/kg Pb, M/F, n = 14-16 (7-8/7-8)
PND 4 to 28
Offspring were
dosed via gavage
every other day
from PND 4 until
PND 28.
PND 29
0.267 |jg/dL for
control
3.27 |jg/dL for
1 mg/kg
12.5 |jg/dL for
10 mg/kg
Offspring Body
Weight
de Fiaueiredo et al.
(2014)
Rat (Wistar)
28 d old Control (untreated), M,
n = 10
60 d old Control (untreated), M, n = 12
28 d old 30 mg/L Pb, M, n = 10
60 d old Control (assumed untreated), M,
n = 12
60 d old 30 mg/L Pb, M, n = 17
PND 0 to PND 28 or
PND 0 to PND 60
Male Wistar rats PND 28
Offspring Body
were dosed via
drinking water
from birth to
PND 28 or 60.
1.2 |jg/dL for control Weight
8.0 |jg/dL 30 mg/L Pb
PND 60
1.6 |jg/dL for control
7.2 |jg/dL for 30 mg/L
Pb
Duan et al. (2017) Mouse (CD-11
PND 1 to PND 21
Dams were dosed
Pups:
Offspring Body
Dams
via drinking water
PND 1
Weight
Control (0 ppm Pb), F, n = 3
starting on GD 1
and continued
1.29 |jg/dL for control
Low dose (27 ppm Pb), F, n = 3
through weaning
(PND 21).
1.29 |jg/dL for low
dose
High dose (109 ppm Pb), F, n = 3
1.29 |jg/dL for high
dose
Pups
PND 18
1.62 |jg/dL for control
Control (0 ppm Pb), NR, n = 9
Low dose (27 ppm Pb), NR, n = 9
19.6 |jg/dL for low
dose
8-184
-------
Exposure
Study Species (Stock/Strain), n, Sex Timing of Exposure BLL Ixamined
Duration)
29.16 |jg/dL for high
High dose (109 ppm Pb), NR, n = 9 dose
PND 35
1.51 |jg/dL for control
28.7 |jg/dL for low
dose
38.0 |jg/dL for high
dose
Betharia and Maher
(2012)
Rat (Sprague-Dawley)
Dams
Control (untreated), F, n = 6
10 |jg/mL Pb, F, n = 6
Pups
Control (untreated), M/F, n = 36-48 (18—
24/18-24)
10 |jg/mL Pb, M/F, n = 36^8 (1824/18-24)
GD Oto PND 20
Dams dosed via
drinking water
starting on GD 0
through weaning
(PND 20).
Pups:
PND 2
0.188 |jg/dL for
control
9.03 |jg/dL for
10 |jg/ml_ Pb
PND 25
0.0880 |jg/dL for
0 |jg/mL
0.976 |jg/dL for
10 |jg/ml_ Pb
Offspring Body
Weight
PND 60
0.0244 |jg/dL for
control
0.0318 |jg/dL for
10 |jg/ml_ Pb
Zhao et al. (2021) Rat (Sprague-Dawley)
Control (untreated), F, n = 6 dams
109 ppm Pb, F, n = 6 dams
GD -14 to PND 10
Dams were dosed Pups:
via drinking water g
starting 2 wk prior „ ,
to mating and 0.87 pg/dL for control
continued until 48.2 |jg/dL for
PND 10. 109 ppm Pb
Offspring Body
Weight
8-185
-------
Exposure
Study Species (Stock/Strain), n, Sex Timing of Exposure BLL Ixamined
Duration)
PND 10
0.87 |jg/dL for control
11.5 |jg/dL for
109 ppm Pb
PND 21
0.87 |jg/dL for control
2.81 |jg/dL for
109 ppm Pb
PND 30
0.87 |jg/dL for control
1.20 |jg/dL for
109 ppm Pb
Rao Barkur and Bairv Rat (Wistar)
(2016) Control (untreated), F, n = 6 dams
0.2% Pb Pregestation Only, n = 6 dams
0.2% Pb Gestation Only, n = 6 dams
0.2% Pb Lactation Only, n = 6 dams
0.2% Pb Gestation and Lactation, F, n = 6
dams
GD -30 to GD -1; GD 0
to GD 21; PND 1 to
PND 21; GD 0 to PND 21
Dams were dosed
via drinking water
for varying
amounts of time:
Pregestation Only
(1 mo prior to
conception),
Gestation Only
(21 d), Lactation
Only (21 d), and
Gestation and
Lactation (42 d).
Pups (PND 22):
0.19 |jg/dL for control
3.03 |jg/dL for 0.2%
Pb in Pregestation
Only group
5.51 |jg/dL for 0.2%
Pb in Gestation Only
group
26.86 |jg/dL for 0.2%
Pb in Lactation Only
group
31.59 |jg/dL for 0.2%
Pb in Gestation and
Lactation group
Offspring Body
Weight, Pinna
Detachment, Eye
Opening, Tooth
Eruption
Barkur and Bairv
(2015)
Rat (Wistar)
Control (untreated),
F, n = 6 dams
GD -30 to GD -1, or
GD 0 to 21, or PND 0 to
21, or GD 0 to PND 21
Dams were dosed
via drinking water
for varying
amounts of time:
Pups (PND 22):
0.18 |jg/dL for control
Offspring Body
Weight
8-186
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Study
Species (Stock/Strain), n, Sex
Timing of Exposure
Exposure
Details
(Concentration,
Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
0.2% Pb Pregestation Only, n = 6 dams
Pregestation Only
(1 mo prior to
0.2% Pb Gestation Only, n = 6 dams
conception),
Gestation Only
0.2% Pb Lactation Only, n = 6 dams
(21 d), Lactation
Only (21 d), and
0.2% Pb Gestation and Lactation, F, n = 6
Gestation and
dams
Lactation (42 d).
3.02 |jg/dL for 0.2%
Pb in Pregestation
Only group
5.30 |jg/dL for 0.2%
Pb Gestation Only
group
26.7 |jg/dL for 0.2%
Pb in Lactation Only
group
32.0 |jg/dL for 0.2%
Pb in Gestation and
Lactation group
Sobolewski et al.
(2020)
Mouse (C57BL/6)
Control (untreated) F, n = 10,
100 ppm Pb, F, n = 10
GD -
only
61 to PND 21 of F1
Dams were dosed
via drinking water
beginning 2 mo
prior to breeding
and ending on
PND 21 of the F1
(weaning).
F1
PND 6-7
0.0 |jg/dL for control,
12.5 |jg/dL for
100 ppm Pb
F3
Postnatal Mo 6-7
0.0 |jg/dL for control,
0.4 |jg/dL for 100 ppm
Pb
Offspring Body
Weight
Albores-Garcia et al.
(2021)
Rat (Long-Evans)
Evaluated on PND 14
Controls (untreated), F, n = 11 dams
Controls (untreated), M/F, n = 14 (7/7) pups
1500 ppm Pb, F, n = 7 dams
1500 ppm Pb, M/F, n = 13 (6/7) pups
Evaluated on PND 28
Controls (untreated), F, n = 9 dams
Continuous exposure
starting at GD -10
Dams were dosed
via the diet
starting 10 d prior
to mating. After
weaning
(PND 21),
offspring were put
onto the same
diet as their
dams.
Pups
PND 14
<1.9 |jg/dL for control
males
<1.9 |jg/dL for control
females
36.1 |jg/dL for
1500 ppm Pb males
37 |jg/dL for
1500 ppm Pb females
Offspring Body
Weight
8-187
-------
Study
Species (Stock/Strain), n, Sex
Timing of Exposure
Exposure
Details
(Concentration,
Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
Controls (untreated), M/F, n = 16 (8/8) pups
1500 ppm Pb, F, n = 8 dams
1500 ppm Pb, M/F, n = 13 (7/6) pups
Evaluated on PND 50
Controls (untreated), F, n = 15 dams
Controls (untreated), M/F, n = 15 (7/8) pups
1500 ppm Pb, F, n = 14 dams
1500 ppm, M/F, n = 15 (7/6) pups
PND 28
<1.9 |jg/dL for control
males
<1.9 |jg/dL for control
females
21.1 |jg/dL for
1500 ppm Pb males
20.9 |jg/dL for
1500 ppm Pb females
Evaluated on PND 120
Controls (untreated), F, n = 13 dams
Control (untreated), M/F, n = 13 (7/6) pups
1500 ppm Pb, F, n = 9 dams
1500 ppm Pb, M/F, n = 12 (6/6) pups
PND 50
<1.9 |jg/dL for control
males
<1.9 |jg/dL for control
females
20.2 |jg/dL for
1500 ppm Pb males
22.1 |jg/dL for
1500 ppm Pb females
PND 120
<1.9 |jg/dL for control
males
<1.9 |jg/dL for control
females
19.6 |jg/dL for
1500 ppm Pb males
24.3 |jg/dL for
1500 ppm Pb females
Basaen and Sobin Mouse (C57BL/6) PND 0 to PND 28 Dams were dosed PND 28 Offspring Body
(2014) Control (untreated), M/F, n = 12 (6/6) via drinking water o.03 |jg/dL for control Weight
from birth of ma|es
8-188
-------
Study
Species (Stock/Strain), n, Sex
Timing of Exposure
Exposure
Details
(Concentration,
Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
30 ppm Pb, M/F, n = 12 (6/6)
330 ppm Pb, M/F, n = 12 (6/6)
offspring until
PND28.
0.03 |jg/dL for control
females
3.63 |jg/dL for 30 ppm
Pb males
2.74 |jg/dL for 30 ppm
Pb females
16.02 |jg/dL for
330 ppm Pb males
13.35 |jg/dL for
330 ppm Pb females
Barkuretal. (2011)
Rat (Wistar)
Control (untreated), F, n = 6 dams
0.2% Pb, F, n = 6 dams
GD 1 to PND 21
Dams were dosed
via drinking water
from GD 1 to
PND 21. Only
male pups were
retained for
measurements of
body weight.
Pups (males only):
PND 22
0.266 |jg/dL for
control
31.2 |jg/dL for 0.2%
Pb
PND 120
0.234 |jg/dL for
control
Offspring Body
Weight
0.468 |jg/dL for 0.2%
Pb
Basha and Reddv
Rat (Wistar)
GD 6 to 21
Dams were dosed
Pups (males only):
Pinna
(2015)
Control (untreated), F, n = 8 dams
via drinking water
PND 21
Detachment,
from GD 6 to
0.21 |jg/dL for control
Tooth Eruption,
Fur
0.2% Pb, F, n = 8 dams
PND 21. Only
male pups were
retained for
measurements of
body weight and
11.2 |jg/dL for 0.2%
Pb
PND 28
Development,
Eye Slit
Formation, Eye
Opening,
8-189
-------
Study
Species (Stock/Strain), n, Sex
Timing of Exposure
Exposure
Details
(Concentration,
Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
developmental
milestones.
0.33 [jg/dL for control
12.3 [jg/dL for 0.2%
Pb
Offspring Body
Weight, Offspring
Body Size
Postnatal Mo 4
0.19 [jg/dL for control
5.9 |jg/dL for 0.2% Pb
BLL = blood lead level; d = day(s); GD = gestational day; F = female; M = male; mo = month(s); NR = not reported; Pb = lead; PND = postnatal day.
8-190
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Table 8-12
Epidemiologic studies of exposure to Pb and puberty in females and puberty in males
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Effects on Puberty in Females
Yao etal. (2019)
United States
2011-2012
Cross-sectional
NHANES
n: 426 female children,
and 470 female
adolescents
Female children (age
6-11 yr) and female
adolescents (age 12-
19 yr) in NHANES
2011-2012
Blood
Blood was measured by ICP-
MS
Age at Measurement:
6-19 yr old
Geometric mean:
Female children: 0.68 |jg/dL
Female adolescents:
0.47 |jg/dL
Median:
Female children: 0.65 |jg/dL
Female adolescents:
0.47 |jg/dL
75th:
Female children: 0.93 |jg/dL
Female adolescents:
0.63 |jg/dL
Quartiles (|jg/dL):
Female children:
Q1: <0.48
Q2: 0.48-0.65
Q3: 0.65-0.93
Q4: >0.93
Female adolescents:
Q1: <0.35
Q2: 0.35-0.47
Puberty among females:
Serum tT levels
Serum tT levels were
analyzed by isotope-
dilution liquid
chromatography-tandem
mass spectrometry
Age at outcome:
6-19 yr old
Weighted multivariable
linear regression
models; Model 1
controlled forage, race,
and BMI. Model 2
controlled for PIR,
seasons of collection,
times of venipuncture,
and serum cotinine, in
addition to the
covariates of model 1
(3 (95% CI), as percent
difference in serum tT
Model 1:
Female children
Q1: Reference
Q2: 14.34 (-3.75, 35.81)
Q3: -5.00 (-21.05, 14.32)
Q4: -5.73 (-23.13, 15.61)
p for trend: 0.36
Female adolescents
Q1
Q2
Q3
Q4
Reference
-8.55 (-18.52, 2.63)
-1.95 (-13.04, 10.56)
13.12 (0.06, 27.88)
p for trend: 0.14
Model 2:
Female children
Q1: Reference
Q2: 14.9 (-3.54, 36.86)
Q3: -0.96 (-17.80, 19.34)
Q4: -2.40 (-21.00, 20.57)
p for trend: 0.63
Female adolescents
Q1
Q2
Q3
Q4
Reference
-7.83 (-18.22, 3.88)
-1.07 (-12.67, 12.06)
14.85 (0.83, 30.81)
p for trend: 0.08
8-191
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Outcome
Confounders
Effect Estimates and 95%
Clsa
Reference and
Study Design
Study Population
Exposure Assessment
Q3: 0.47-0.63
Q4: >0.63
Siawiriska et al.
(2012)
Legnica-Glogow
District
Poland
1995-2007
Cross-sectional
1995 n:436; 2007
n:346
Menarche status of
schoolgirls 7-16 yr
from villages in
southwestern Poland
was surveyed in 1995,
2001, 2004, and 2007.
Blood
Blood was measured by
GFAAS with a Zeeman
correction for background
Age at Measurement:
7-16 yr old
Mean
1995: 6.57 pg/dL
2007: 4.24 pg/dL
Puberty among females:
Short-term secular
change in menarche
Menarche through survey
Age at outcome:
7-16 yr
Logistic regression
models were adjusted
for age, height (linear
growth), BMI (weight-
for-height), and Pb
group (low Pb group:
2-5 pg/dL; high Pb
group: 5.10-
33.90 pg/dL)
OR (95% CI)
1995: 0.70 (0.27,
2007: 0.31 (0.09,
1.85)
1.06)
OR (95% CI)
Model with BMI: 0.54 (0.26,
1.13)
Model with percent body fat:
0.52 (0.25, 1.08)
Model with sum of skinfolds:
0.53 (0.26, 1.10)
Cross-sectional 010
Mean
Total: 3.6 pg/dL
<3.7 pg/dL: 2.9 pg/dL
>3.7 pg/dL: 4.4 pg/dL
Median
Total: 3.6 pg/dL
<3.7 pg/dL: 2.8 pg/dL
>3.7 pg/dL: 4.3 pg/dL
Gomula et al. (2022) n: 490
Polkowice
Poland
2008
Girls aged 7-16 yr
who were attending
several schools in
Polkowice in 2008.
Blood
Blood was measured by AAS
with Zeeman background
correction
Age at measurement: 7-16 yr
Puberty among females:
age at menarche
Menarche through survey
Age at outcome: 7-16 yr
old
Logistic regression
models were adjusted
for age and (1) BMI; (2)
percent body fat; and
(3) sum of skinfolds
8-192
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
De Craemer et al.
(2017)
Belgium
FLEHS I: 2002-2006,
FLEHS II: 2007-
2011, and FLEHS III:
2012-2015
Cross-sectional
FLEHS I, FLEHS II
and FLEHS III
n: FLEHS I: n = 1659,
FLEHS II: n = 606, and
FLEHS III: n = 406
Adolescents aged 14-
15 yr
Blood
Blood Pb was measured by
ICP-MS
Age at Measurement:
14-15 yr old
Geometric meanb
FLEHS I: 2.13 pg/dL
FLEHS II: 1.38 pg/dL
FLEHS III: 0.926 pg/dL
Maxb
FLEHS I: 21.2 pg/dL
FLEHS II: 7.69 pg/dL
FLEHS III: 3.86 pg/dL
Puberty among females:
Hormones and sexual
maturation in adolescents
Development of breasts in
adolescent females and
pubic hair was scored
using the international
scoring criteria of
Marshall and Tanner,
where stage 1
corresponds to the start of
puberty and stage 5 to the
adult stage. Information
on menarche was
obtained through self-
assessed questionnaires.
Age at outcome:
14-15 yr old
Logistic regression
models for female pubic
hair development and
breast development
were adjusted for age
BMI, contraceptive pill
usage; linear
regression models for
age at menarche were
adjusted for age, BMI
OR (95% Cl)c
Breast development
FLEHS I: 0.798 (0.653,
0.969)
FLEHS II: 1.318 (0.936,
2.055)
FLEHS III: 1.187 (0.886,
1.627)
Pubic hair development
FLEHS I: 1.113 (0.922,
1.349)
FLEHS II: 1.322 (0.938,
2.083)
FLEHS III: 0.919 (0.677,
1.229)
(3 (95% Cl)c
Age of menarche
FLEHS I: 0.039 (-0.072,
0.15)
FLEHS II: 0.257 (0.091,
0.424)
FLEHS III: 0.126 (-0.021,
0.273)
Nkomo et al. (2018)
Johannesburg
South Africa
Cohort
BT20+ birth cohort
n: 683
Singleton births in
which the infant
resides in
Johannesburg area for
at least 6 mo after
birth; participants must
have data for BLL at
age 13 and pubertal
Blood and cord blood
UCB collected at birth and
blood at collected at age 13
were measured by AAS with
a Zeeman background
correction
Age at Measurement:
birth and age 13
Puberty among females:
Pubertal trajectory
classes
Tanner stages of pubertal
development refer to a
standard clinical method
used to describe physical
measurements of
secondary sexual
characteristics using
Multinomial logistic
regression was used to
predict pubertal growth
trajectory class based
on BLLs at age 13 yr
and cord BLLs adjusted
for ethnicity and height
at age 8
RR (95% CI)
Development of pubic hair
UCB
Blood, >5 pg/dL vs.
<5 pg/dL
Trajectory Class 1:
Reference
Trajectory Class 2: 0.45
(0.29, 0.68)
8-193
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
growth trajectory
classes
Mean (SD)
UCB: 5.8 (2.1) pg/dL
Blood: 5.0 (1.9) pg/dL
Median
UCB: 6.0 |jg/dL
Blood: 4.8 |jg/dL
75th
UCB: 7.0 |jg/dL
Blood: 7.9 |jg/dL
drawings to signal stage
of pubertal development
where stage 1 signifies
lowest level of pubertal
maturation and stage 5
denotes highest level of
pubertal maturation in
girls
Age at outcome:
9-16 yr old
Trajectory Class 3: 0.55
(0.26, 1.17)
Development of breasts
Blood, >5 pg/dL vs.
<5 pg/dL
Trajectory Class 1:
Reference
Trajectory Class 2: 0.72
(0.47, 1.11)
Trajectory Class 3: 0.63
(0.42, 0.94)
Trajectory Class 4: 0.46
(0.27, 0.77)
Liu etal. (2019b)
Mexico City
Mexico
Cohort
n: 547 (283 girls and
264 boys)
Pregnant women were
recruited at three
public maternity
hospitals (Manuel Gea
Gonzalez Hospital,
Mexican Social
Security Institute and
the National Institute
of Perinatology) in
Mexico City; and
Children at age 9.8-
18.0 yr who had at
least one
measurement of
maternal bone Pb or
childhood blood Pb
Blood and bone
Maternal bone, measured at
the mid-tibial shaft (cortical
bone) and patella (trabecular
bone) was measured by K-
XRF instrument; blood
samples from children were
measured by GFAAS
Age at Measurement:
Maternal age 1-mo
postpartum; blood measured
between 1 and 4 yr
Median
Patella: 8.20 pg/g
Tibia: 7.63 pg/g
Blood, cumulative 1-4 yr:
13.83 pg/dL
75th
Puberty among females:
Pubertal stages
In girls, the stages of
pubertal development
were defined by a
pediatrician using Tanner
staging scales for the
breast maturation and
pubic hair growth.
Menarche was measured
via a self-reported
questionnaire.
Age at outcome:
9.8-18 yr
Ordinal regression
models were adjusted
for child age at visit,
maternal education and
marital status, and
number of siblings at
birth; Cox proportional
hazard regression
models were adjusted
for number of siblings
at birth, maternal
education, and marital
status
OR (95% CI), per IQR
increase in Pb
Breast development
Patella: 0.79 (0.61, 1.01)
Tibia: 1.01 (0.75, 1.36)
Blood, cumulative 1-4 yr:
0.96 (0.92, 0.99)
Pubic hair development
Patella: 0.96 (0.76, 1.22)
Tibia: 1.12 (0.84, 1.49)
Blood, cumulative 1-4 yr:
0.95 (0.92, 0.99)
HR (95% CI)
Patella
Continuous: 0.16 (0.02,
1.07)
T1: Reference
T2: 1.10 (0.76, 1.58)
8-194
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Outcome
Confounders
Effect Estimates and 95%
Clsa
T3: 0.60 (0.41, 0.88)
Tibia
Continuous: 1.11 (0.12,
9.84)
T1: Reference
T2: 1.30 (0.86, 1.96)
T3: 1.14 (0.75, 1.72)
Blood, cumulative 1-4 yr
Continuous: 0.91 (0.77,
1.08)
T1
T2
T3
Reference
0.65 (0.46, 0.91)
0.76 (0.55, 1.06)
Reference and
Study Design
Study Population
Exposure Assessment
Patella: 15.45 |jg/g
Tibia: 13.80 |jg/g
Blood, cumulative 1-4 yr:
18.76 |jg/dL
IQR
Patella: 13.57 |jg/g
Tibia: 13.30 |jg/g
Blood, cumulative 1-4 yr:
7.66 |jg/dL
Tertiles
Patella (pg/g)
T1
T2
T3
<3.9
4.0-12.9
13.0-45.3
Tibia (|jg/g)
T1: <4.6
T2: 4.7-11.3
T3: 11.4-37.3
Blood, cumulative 1-4 yr
(Hg/dL)
T1
T2
T3
<12.0
12.1-16.1
16.2-51.5
Jansen et al. (2018) ELEMENT project
Mexico City
Mexico
1997-2004(2015)
Cohort
n: 200
Mothers were
recruited from prenatal
clinics ofthe Mexican
Social Security
Institute in Mexico City
who were not planning
to leave the area
Blood
Maternal blood was
measured by GFAAS
Age at Measurement:
maternal age at sampling
Median
Puberty among females:
Menarche
Girls were asked about
menarche during the
follow-up visit (between
age 9.8 and 18.1 yr).
They were asked whether
or not menarche had
occurred (Yes, no, or
Interval-censored Cox
regression models,
comparing the hazard
of menarche among
girls with prenatal
maternal blood Pb
>5 |jg/dL to those with
prenatal maternal BLL
<5 |jg/dL, were
adjusted for maternal
HR (95% CI)
Interval-censored Cox
models
First trimester maternal
blood
<5 |jg/dL: Reference
>5 |jg/dL: 0.85 (0.46, 1.24)
8-195
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
within 5 yr; had a
history of infertility,
diabetes, or psychosis;
consuming alcoholic
beverages daily during
pregnancy; addiction
to illegal drugs;
diagnosis of a high-
risk pregnancy; or
being pregnant with
multiples
First trimester: 4.8 [jg/dL
Second trimester: 4.0 [jg/dL
Third trimester: 4.5 [jg/dL
75th:
First trimester: 7.1 [jg/dL
Second trimester: 6.4 [jg/dL
Third trimester: 6.6 [jg/dL
don't know/refused) and,
if so, to recall the age (in
years and months) it
occurred.
Age at outcome:
age of menarche
age, maternal parity,
maternal education,
and prenatal calcium
treatment status; Cox
regression models,
using self-reported age
at menarche as the
time to event, were
adjusted for maternal
age, maternal parity,
maternal education,
and prenatal calcium
treatment status; Cox
regression models were
also restricted to girls
<14.5 yr at the time of
the interview and
adjusted for maternal
age, maternal parity,
maternal education,
and prenatal calcium
treatment status
Second trimester maternal
blood
<5 [jg/dL: Reference
>5 [jg/dL: 0.59 (0.28, 0.90)
Third trimester maternal
blood
<5 [jg/dL: Reference
>5 [jg/dL: 0.85 (0.42, 1.27)
Cox models
First trimester maternal
blood
<5 [jg/dL: Reference
>5 [jg/dL: 0.92 (0.65, 1.29)
Second trimester maternal
blood
<5 [jg/dL: Reference
>5 [jg/dL: 0.91 (0.65, 1.27)
Third trimester maternal
blood
<5 [jg/dL: Reference
>5 [jg/dL: 0.97 (0.69, 1.37)
Cox models restricted to
girls <14.5 yr at interview
First trimester maternal
blood
<5 [jg/dL: Reference
>5 [jg/dL: 0.80 (0.52, 1.25)
Second trimester maternal
blood
<5 [jg/dL: Reference
>5 [jg/dL: 0.64 (0.38, 1.09)
8-196
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Third trimester maternal
blood
<5 |jg/dL: Reference
>5 |jg/dL: 0.89 (0.56, 1.41)
Effects on Puberty Among Males
Yao etal. (2019)
United States
2011-2012
Cross-sectional
NHANES
n: 431 male children,
493 male adolescents
Male children (age 6-
11 yr) and male
adolescents (age 12-
19 yr) in NHANES
2011-2012
Blood
Blood was measure by ICP-
MS
Age at Measurement:
6-19 yr old
Geometric mean
Male children: 0.76 |jg/dL
Male adolescents: 0.68 |jg/dL
Median
Male children: 0.72 |jg/dL
Male adolescent: 0.66 |jg/dL
75th
Male children: 1.02 |jg/dL
Male adolescents: 0.96 |jg/dL
Quartiles (|jg/dL):
Male children:
Q1
Q2
Q3
Q4
<0.52
0.52-0.72
0.72-1.02
>1.02
Male adolescents:
Q1: <0.47
Puberty among males:
Serum tT levels in male
children and adolescents
Serum tT levels were
analyzed by isotope-
dilution liquid
chromatography-tandem
mass spectrometry
Age at outcome:
6-19 yr old
Weighted multivariable
linear regression
models; Model 1
controlled forage, race,
and BMI. Model 2
controlled for PIR,
seasons of collection,
times of venipuncture,
and serum cotinine, in
addition to the
covariates of model 1
(3 (95% CI), as percent
difference in serum tT
Model 1:
Male children
Q1: Reference
Q2: 4.1 (-18.47, 32.9)
Q3: -6.13 (-27.64, 21.77)
Q4: -12.83 (-33.68, 14.58)
p for trend: 0.36
Male adolescents
Q1
Q2
Q3
Q4
Reference
-3.36 (-20.98, 18.2)
14.99 (-7.77, 43.37)
15.62 (-7.07, 43.86)
p for trend: 0.18
Model 2:
Male children
Q1: Reference
Q2: 11.75 (-13.06, 43.65)
Q3: -4.63 (-26.97, 24.55)
Q4: -13.09 (-34.45, 15.22)
p for trend: 0.42
Male adolescents
Q1: Reference
Q2: -4.35 (-21.22, 16.14)
8-197
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Q2: 0.47-0.66
Q3: 0.66-0.96
Q4: >0.96
Q3: 8.15 (-12.91, 34.3)
Q4: 6.32 (-14.62, 32.4)
p for trend: 0.58
De Craemer et al.
(2017)
Belgium
FLEHS I: 2002-2006,
FLEHS II: 2007-
2011, and FLEHS III:
2012-2015
Cross-sectional
FLEHS I, FLEHS II
and FLEHS III
FLEHS I n: 1659,
FLEHS II n: 606, and
FLEHS III n: 406
Adolescents aged 14-
15 yr
Blood
Blood was analyzed by ICP-
MS
Age at Measurement:
14-15 yr old
Geometric meanb
FLEHS I: 2.13 pg/dL
FLEHS II: 1.38 pg/dL
FLEHS III: 0.926 pg/dL
Maxb:
FLEHS I: 21.2 pg/dL
FLEHS II: 7.69 pg/dL
FLEHS III: 3.86 pg/dL
Puberty among males:
Hormones and sexual
maturation in adolescents
Development of genitals
in adolescent males and
pubic hair was scored
using the international
scoring criteria of
Marshall and Tanner,
where stage 1
corresponds to the start of
puberty and stage 5 to the
adult stage. Sex
hormones investigated in
this study were E2,
testosterone, fE2 and fT,
SHBG, LH, and FSH.
Hormone levels in
adolescent males were
measured in blood serum
using commercial
immunoassays.
Age at outcome:
14-15 yr old
Logistic regression
models for male public
hair development and
genital development
were adjusted for age
and BMI; linear
regression models for
hormones (ratio T/E2,
E2, fE2, T, IT) were
adjusted for age, hr of
blood collection, BMI,
smoking status; SHBG:
age, fasting, BMI,
smoking status, hr of
blood collection; LH
and FSH: age, BMI,
smoking status
OR (95% Cl)c
Pubic hair development
FLEHS I: 0.808 (0.686,
0.949)
FLEHS II: 0.849 (0.563,
1.365)
FLEHS III: 0.515 (0.327,
0.774)
Genital development
FLEHS I: 0.843 (0.717,
0.99)
FLEHS II: 0.697 (0.462,
0.998)
FLEHS III: 0.621 (0.388,
0.967)
(3 (95% Cl)c
FLEHS I:
Ratio T/E2: 1.022 (0.985,
1.059)
E2: 1.011 (0.991, 1.031)
fE2: 1.003 (0.975, 1.033)
T: 1.039 (0.993, 1.087)
fT: 1.026 (0.967, 1.09)
SHBG: 1.024 (0.992, 1.056)
LH: 0.995 (0.959, 1.033)
FLEHS II:
ratio T/E2: 1.002 (0.958,
1.049)
8-198
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
E2: 0.968 (0.923, 1.016)
fE2: 0.908 (0.839, 0.983)
T: 0.959 (0.906, 1.015)
IT: 0.909 (0.828, 0.997)
SHBG: 1.005 (0.961, 1.052)
LH: 0.974 (0.923, 1.028)
FSH: 0.995 (0.942, 1.05)
Nkomo et al. (2018)
Johannesburg
South Africa
Cohort
BT20+ birth cohort
n: 683
Singleton births in
which the infant
resides in
Johannesburg area for
at least 6 mo after
birth; participants must
have data for BLL at
age 13 and pubertal
growth trajectory
classes
Blood and cord blood
UCB collected at birth and
blood at collected at age 13
were measured by AAS with
a Zeeman background
correction
Age at Measurement:
birth and age 13
Mean (SD)
UCB: 5.9 (2.0) pg/dL
Blood: 6.6 (2.6) pg/dL
Median
UCB: 6.0 pg/dL
Blood: 6.5 pg/dL
75th
UCB: 7.0 pg/dL
Blood: 6.0 pg/dL
Puberty among males:
Pubertal trajectory
classes
Tanner stages of pubertal
development refer to a
standard clinical method
used to describe physical
measurements of
secondary sexual
characteristics using
drawings to signal stage
of pubertal development
where stage 1 signifies
lowest level of pubertal
maturation and stage 5
denotes highest level of
pubertal maturation in
boys
Age at outcome:
9-16 yr old
Multinomial logistic
regression models were
used to predict pubertal
growth trajectory class
based on (1) UCB Pb
and adjusted for
ethnicity; (2) blood Pb
and adjusted for
ethnicity and height at
age 8
RR (95% CI)
UCB
Pubic hair development
Trajectory Class 1:
Reference
Trajectory Class 2: 0.61
(0.25, 1.43)
Trajectory Class 3: 0.28
(0.11, 0.74)
Genital development
Trajectory Class 1:
Reference
Trajectory Class 2: 0.27
(0.03, 2.26)
Trajectory Class 3: 0.24
(0.03, 1.89)
Trajectory Class 4: 0.13
(0.01, 1.24)
Blood
Pubic hair development
Trajectory Class 1:
Reference
Trajectory Class 2: 0.94
(0.63, 1.39)
8-199
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Trajectory Class 3: 1.35
(0.73, 2.47)
Genital development
Trajectory Class 1:
Reference
Trajectory Class 2: 0.77
(0.33, 1.77)
Trajectory Class 3: 0.88
(0.38, 2.01)
Trajectory Class 4: 1.02
(0.37, 2.83)
Liu etal. (2019b)
Mexico City
Mexico
Cohort
n: 547 (283 girls and
264 boys)
Pregnant women were
recruited at three
public maternity
hospitals (Manuel Gea
Gonzalez Hospital,
Mexican Social
Security Institute and
the National Institute
of Perinatology) in
Mexico City; and
Children at age 9.8-
18.0 yr who had at
least one
measurement of
maternal bone Pb or
childhood blood Pb
Blood and bone
Maternal bone was measured
at the mid-tibial shaft (cortical
bone) and patella (trabecular
bone) and determined using
the X-ray fluorescence
instrument; blood samples
from children were measured
by GFAAS
Age at Measurement:
Maternal age 1-mo
postpartum; blood measured
between 1 and 4 yr
Median
Patella: 7.44 |jg/g
Tibia: 7.10 |jg/g
Blood, cumulative 1-4 yr:
14.33 |jg/dL
75th
Patella: 14.56 |jg/g
Puberty among males:
Pubertal stages
In boys, the stage of
sexual maturation was
defined by the
pediatrician using Tanner
staging scales for the
development of genitalia
and pubic hair.
Age at outcome:
9.8-18 yr
Ordinal regression
models for genitalia
and pubic hair and
logistic regression
models for TV were
adjusted for adjusted
for child age at visit,
maternal education and
marital status, and
number of siblings at
birth
OR (95%), per IQR increase
in Pb
Genital development
Patella: 0.963 (0.734,
1.264)
Tibia: 1.00 (0.711, 1.406)
Blood, cumulative 1-4 yr:
0.995 (0.948, 1.044)
Pubic hair development
Patella: 1.094 (0.836,
1.432)
Tibia: 1.00 (0.715, 1.398)
Blood, cumulative 1-4 yr:
1.004 (0.969, 1.04)
TV
Patella: 1.158 (0.804,
1.667)
Tibia: 0.885 (0.503, 1.558)
Blood, cumulative 1-4 yr:
1.013 (0.954, 1.075)
8-200
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Confounders
Effect Estimates and 95%
Clsa
Reference and
Study Design
Study Population
Exposure Assessment
Tibia: 15.93 |jg/g
Blood, cumulative 1-4 yr:
18.90 |jg/dL
IQR
Patella: 13.57 |jg/g
Tibia: 13.30 |jg/g
Blood, cumulative 1-4 yr:
7.66 |jg/dL
Williams etal. (2019)
Chapaevsk
Russian
2003-2005(2017)
Cohort
Russian Children's
Study
n: 516
Healthy male children
who were 8-9 yr old
between 2003 and
2005 in Chapaevsk,
Russia.
Blood
Blood was measured by
Zeeman background
corrected flameless GFAAS
Age at Measurement:
8-9 yr old
Median: 3 |jg/dL
Max: 31 |jg/dL
Puberty among males:
Male sexual maturity
Pubertal status was
staged from 1 to 5 via
examination by a single
clinician according to
internationally accepted
criteria. Pubarche (pubic
hair stage, P) was
determined by the extent
of terminal hair growth.
Genital staging (G) was
assessed by genital size
and maturity. TV was
measured using an
orchidometer. Three
different measures of
sexual maturity were
considered as separate
indicators: TV >20 mL of
either testis, genitalia
stage 5 (G5), and pubic
hair stage 5 (P5).
Duration of pubertal
progression was defined
as time from pubertal
onset (TV >3 mL,
genitalia stage >2 (G2),
Interval-censored
models were fit
assuming a normal
distribution for age at
sexual maturity using
accelerated failure time
models to compare
pubertal outcomes
between boys with
'higher' (>5 |jg/dL)
versus 'lower'
(<5 |jg/dL) peripubertal
BLLs. Models were
adjusted for boy's BW,
prenatal exposure to
maternal alcohol and
tobacco, maternal age
at son's birth,
household
characteristics including
income level, parental
education, and whether
the biological father
lived in the same
household, the boy's
physical activity, and
his nutritional status
determined by caloric
(3 (95% Cl)c, as shift in
mean age in months
Age at pubertal onset
Genitalia (G2): 8.40 (3.70,
13.10)
Pubic hair (P2): 8.12 (3.46,
12.78)
TV (>3 mL): 7.68 (3.46,
11.90)
Age at sexual maturity
Genitalia (G5): 4.20 (0.56,
7.84)
Pubic hair (P5): 4.23 -0.31,
8.77)
TV (>20 mL): 5.14 (1.70,
8.58)
Duration of puberal
progression
Genitalia (G2 to G5): -3.76
(-7.93, 0.42)
Pubic hair (P2 to P5): -1.82
(-6.91, 3.28)
TV (>3 mL to >20 mL):
-1.19 (-4.92, 2.54)
8-201
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Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
pubic hair stage >2 (P2),
respectively) to sexual
maturity, separately for
each pubertal indicator.
Age at outcome:
age at follow-up in 2017
intake and percent of
fat and protein intake.
Mediation analysis was
conducted to partition
the effect of higher vs.
lower BLLs on the age
at sexual maturity into a
direct effect of Pb
exposure and indirect
effect of Pb acting
through HTZ and BMIZ
(mediators) at age 11.
Mediation Analysis, as % of
total
HTZ
G5: 53.0% ((3: 2.37 mo)
P5: 47.5% ((3: 2.36 mo)
TV >20 mL: 34.2% ((3:
1.78 mo)
BMIZ
G5: 14.3% ((3: 0.64 mo)
P5: 23.4% ((3: 1.16 mo)
TV >20 mL: 6.1% ((3:
0.32 mo)
Fleisch et al. (2013)
Chapaevsk
Russia
2003-2005
Follow-up: 2-yr (at
10-11 yr) and 4-yr (at
12-13 yr)
Cohort
Russian Children's
Study
n: 394
Boys ages 8-9 yr old
from Chapaevsk,
Russia
Blood
Blood was measured by
Zeeman background
corrected flameless GFAAS
Age at Measurement:
8-9 yr old
Median: 3 |jg/dL
75th: 5 |jg/dL
Max: 31 |jg/dL
Puberty among males:
IGF-1
Serum IGF-1
concentrations were
measured by a
chemiluminescent
immunometric assay
using Siemens Immulite
2000.
Age at outcome:
10-11 yr (at 2-yr follow-
up); 12-13 (at 4-yr follow-
up)
Linear regression
models using a GEE
approach to account for
the repeated measures
were fitted to predict
the mean levels of
serum concentrations
of IGF-1 (ng/mL) in
relation to BLLs,
adjusted for baseline
parental education,
BW, nutritional intake,
and baseline and
follow-up age and BMI
(3 (95% Cl)c, as adjusted
mean change
BLL <5 |jg/dL: Reference
BLL >5 ug/dL: -29.2 ng/mL
(-43.8, -14.5)
Pre-puberty
BLL <5 |jg/dL: Reference
BLL >5 ug/dL: -14.1 ng/mL
(-0.9, -27.2)
Early puberty:
BLL <5 |jg/dL: Reference
BLL >5 |jg/dL: -18.0 (-3.5,
-32.5)
Mid-puberty
BLL <5 |jg/dL: Reference
BLL >5 ug/dL: -41.9 ng/mL
(-15.1, -68.7)
8-202
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Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
AAS = atomic absorption spectrometry; BMI = body mass index; BMIZ = BMI-for-age Z-score; BT20+ = Birth to Twenty Plus; BW = birth weight; E2 = estradiol; ELEMENT = Early
Life Exposure in Mexico to Environmental Toxicants; fE2 = free estradiol; FLEHS = Flemish Environment and Health Study; FSH = follicle stimulating hormone; fT = free
testosterone; GEE = generalized estimating equation; GFAAS = graphite furnace atomic absorption spectrometry; HR = hazard ratio; HTZ = height Z-score; ICP-MS = inductively
coupled plasma mass spectrometry; IGF-1 = insulin-like growth factor 1; LH = luteinizing hormone; mo = month(s); NHANES = National Health and Nutrition Examination Survey;
OR = odds ratio; PIR = poverty-income ratio; RR = relative risk; SD = standard deviation; SHBG = sex hormone binding globulin; T = testosterone; tT = total testosterone;
TV = testicular volume; UCB = umbilical cord blood; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bPb measurements were converted from |jg/L to |jg/dL.
°Effect estimates unable to be standardized.
8-203
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Table 8-13 Epidemiologic studies of exposure to Pb and other developmental effects
study Desfgn Study P°Pulation ExP°sure Assessment Outcome Confounders Effect and 95%
Alearia-Torres et al.
(2020)
Salamanca
Mexico
Cross-sectional
n: 86
Healthy children 6-
15 yr of age were
recruited from four
primary schools
Blood
Blood was measured by
ICP-MS
Age at Measurement:
6-15 yr old
Mean (SD): 3.78
(3.73) |jg/dL
Max: 22.61 pg/dL
Other developmental
effects: Telomeric
lengthening and mtDNA
effects
DNA was isolated from
peripheral blood and rTL
and the mtDNAcn were
determined by real-time
polymerase chain reaction
Age at outcome:
6-15 yr old
Linear regression
analyses; TL models
were adjusted for
mtDNAcn, sex, age, and
total white blood cell
count; mtDNAcn models
adjusted for TL, sex,
age, and total white
blood cell count
(3 (95% Cl)b
TL: 0.088 (-0.027, 0.097)
mtDNAcn: -0.198 (-2.81,
-0.17)
Tamavo v Ortiz et al.
(2016)
Mexico City
Mexico
2007-2011
Cohort
PROGRESS birth
cohort
n: 255 for 12 mo
n: 150 for 18-24 mo
Women were invited to
participate during their
prenatal care visits at 4
clinics belonging to the
Mexican Social Security
System
Blood and bone
Maternal blood, collected
twice during pregnancy
(second and third
trimesters), was
measured by ICP-MS.
Maternal bone, from the
mid-tibial shaft, was
measured using a K-XRF
instrument during the first
month postpartum visit
Age at Measurement:
Maternal age at second
trimester, third trimester,
and 1 mo postpartum
Mean
2nd trimester blood for 12-
mo-old infants: 3.5 pg/dL
Other developmental
effects: Cortisol levels
Four saliva samples per
day from their child at
home; saliva samples
were analyzed in
duplicate using a
chemiluminescence-assay
Age at outcome:
12 or 18-24 mo
Longitudinal functional
mixed effects regression
models with penalized
splines were adjusted for
child's sex and maternal
age at delivery,
education, and pre-
pregnancy BMI
(3 (95% Cl)b
12-mo infants
Second trimester maternal
blood
Lower Pb: Reference
Moderate Pb: -0.07 (-0.24,
0.10)
Higher Pb: -0.51 (-0.85,
-0.18)
Third trimester maternal
blood
Lower Pb: Reference
Moderate Pb: -0.14 (-0.31,
0.03)
Higher Pb: -0.02 (-0.31,
0.26)
Tibia
Lower Pb: Reference
8-204
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
2nd trimester blood for
18-24-mo-old infants:
3.9 |jg/dL
3rd trimester blood for 12-
mo-old infants: 3.7 |jg/dL
3rd trimester blood for 18-
24-mo-old infants:
4.2 |jg/dL
Tibia for 12-mo-old
infants: 5.6 |jg/g
Tibia for 18-24-mo-old
infants: 4.9 |jg/g
Tertiles
Lower Pb: <5 |jg/dL
Moderate Pb: 5 < Pb
<10 |jg/dL
High Pb: >10 |jg/dL
Moderate Pb: 0.02 (-0.14,
0.19)
Higher Pb: -0.03 (-0.21,
0.14)
18-24-mo infants
Second trimester maternal
blood
Lower Pb: Reference
Moderate Pb: 0.11 (-0.08,
0.30)
Higher Pb: 0.23 (-0.19, 0.65)
Third trimester maternal
blood
Lower Pb: Reference
Moderate Pb: 0.01 (-0.17,
0.20)
Higher Pb: -0.05 (-0.51,
0.41)
Tibia
Lower Pb: Reference
Moderate Pb: 0.10 (-0.13,
0.32)
Higher Pb: 0.14 (-0.08, 0.35)
Hou et al. (2020)
Guiyu and Haojiang
China
November-
December 2017
Cross-sectional
n: 574 (357 from Guiyu
and 217 from Haojiang)
Children 2.5-6 yr of
age that lived in Guiyu,
an e-waste
contaminated town or
Haojiang, a city with
Blood
Blood was measured by
GFAAS
Age at Measurement:
2.5-6 yr old
Median
Other developmental
effects: Oral anti-
inflammatory potential
Participants were
instructed to sit up straight
and slightly forward in
their chair. A sputum cup
was used to collect the
saliva. Decayed
Multivariable linear
regression model
adjusted for gender, age,
BMI, outdoor activities,
the sucking/biting of toys
and pencils, diet (sweet
consumption, bean
products, marine
products), family
member smoking,
(3 (95% Cl)b: -3.65 (-8.07,
0.77)
8-205
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Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
similar culture but no e-
waste recycling activity
Reference group:
3.47 |jg/dL
Exposed group:
4.86 |jg/dL
75th
Reference group:
4.07 |jg/dL
Exposed group:
4.86 |jg/dL
deciduous teeth were
detected under natural
and artificial light. The
concentration of salivary
sialic acids was
determined using a
quantitative competitive
ELISA kit.
Age at outcome:
2.5-6 yr old
paternal education
levels, monthly
household income
Sitarik et al. (2020)
Detroit, Ml
United States
September 2003-
December 2007
(December 2011-
September 2019)
Cohort
WHEALS birth cohort
n: 146
All women were in their
second trimester or
later, were aged 21-
49 yr, and were living in
a predefined
geographic area in
Wayne and Oakland
counties of Michigan.
Teeth were selected for
metal measurement if
(1) the child had at
least some outcome
data available (birth
outcomes and/or a 2-yr
clinic visit) or early life
microbiome data; and
(2) the tooth sample
met laboratory quality
control/quality
assurance guidelines
Teeth
Teeth were measured by
LA-ICP-MS. Teeth were
sectioned, and the
neonatal line (a
histological feature formed
in enamel and dentine at
the time of birth) and
incremental markings
were used to assign
temporal information to
sampling points. Second
trimester, third trimester,
postnatal (birth through 1
yr), and childhood (age 1
to tooth shedding) Pb
levels.
Age at measurement:
Estimated exposure from
2nd trimester, 3rd
trimester, and postnatally
(<1 yr of age)
Other developmental
effects: Gut microbiota (in
infants)
Families were asked to
retain the most recent
soiled diaper prior to the
home visit and stool
samples from infants ages
1-6 mo.
Age at outcome:
1-6 mo
Permutational
multivariate analysis of
variance models were
adjusted for tooth type,
tooth attrition, tooth
batch, exact age at stool
sample collection, and
child race
(3 (SE)b
Alpha diversity metrics
Second trimester
Richness - Bacterial
1 mo: 5.53 (6.98)
6 mo: -7.77 (7.31)
Richness - Fungal
1 mo: 0.29 (1.65)
6 mo: 1.7 (1.51)
Evenness - Bacterial
1 mo: 0 (0.01)
6 mo: -0.02 (-0.01)
Evenness - Fungal
1 mo: 0.03 (0.05)
6 mo: -0.02 (0.05)
Faith's Diversity - Bacterial
1 mo: 0.16 (0.39)
6 mo: -0.19 (0.37)
Faith's Diversity - Fungal
1 mo: Not reported
6 mo: Not reported
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Shannon Diversity -
Bacterial
1 mo: 0.01 (0.08)
6 mo: -0.11 (0.07)
Shannon Diversity - Fungal
1 mo: 0.06 (0.15)
6 mo: 0 (0.14)
Third trimester
Richness - Bacterial
1 mo: 2.52 (6.37)
6 mo: -13.11 (8.36)
Richness - Fungal
1 mo: 0.69 (1.82)
6 mo: 2.54 (1.56)
Evenness - Bacterial
1 mo: -0.01 (0.01)
6 mo: -0.02 (-0.01)
Evenness - Fungal
1 mo: 0.03 (0.05)
6 mo: 0.03 (0.05)
Faith's Diversity - Bacterial
1 mo: 0.03 (0.35)
6 mo: -0.52 (0.42)
Faith's Diversity - Fungal
1 mo: Not reported
6 mo: Not reported
Shannon Diversity -
Bacterial
1 mo: -0.05 (0.07)
6 mo: -0.12 (0.08)
Shannon Diversity - Fungal
1 mo: 0.09 (0.16)
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Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
6 mo: 0.15 (0.15)
Postnatal
Richness - Bacterial
1 mo: 2.18 (7.16)
6 mo: -2.55 (6.42)
Richness - Fungal
1 mo: -1.85 (2.54)
6 mo: -0.35 (1.05)
Evenness - Bacterial
1 mo: -0.02 (0.01)
6 mo: -0.01 (0.01)
Evenness - Fungal
1 mo: 0.07 (0.1)
6 mo: 0.06 (0.06)
Faith's Diversity - Bacterial
1 mo: -0.08 (0.39)
6 mo: 0.11 (0.32)
Faith's Diversity - Fungal
1 mo: Not reported
6 mo: Not reported
Shannon Diversity -
Bacterial
1 mo: -0.1 (-0.08)
6 mo: -0.05 (-0.06)
Shannon Diversity - Fungal
1 mo: -0.07 (0.23)
6 mo: -0.05 (0.1)
BMI = body mass index; ELISA = enzyme-linked immunosorbent assay; GFAAS = graphite furnace atomic absorption spectrometry; ICP-MS = inductively coupled plasma mass
spectrometry; K-XRF = K-shell X-ray fluorescence instrument; LA-ICP-MS = laser ablation-inductively coupled plasma-mass spectrometry; mo = month(s); mtDNA = mitochondrial
DNA; mtDNAcn = mitochondrial DNA copy number; rTL = relative telomere length; SD = standard deviation; SE = standard error; TL = telomere length; WHEALS = Wayne County
Health, Environment, Allergy and Asthma Longitudinal Study; yr = year(s).
8-208
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study Desfgn Study P°Pulation ExP°sure Assessment Outcome Confounders Effect Esti™£s and 95%
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bEffect estimates unable to be standardized.
8-209
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Table 8-14
Epidemiologic studies of exposure to Pb and female reproductive effects
Reference and
Study Design
Study Population Exposure Assessment Outcome Confounders
Effect Estimates and 95%
Clsa
Effects on Hormones Levels and Menstrual/Estrous Cycle
Krieq and Feng
(2011)
United States
1999-2002
Cross-sectional
NHANES
n: 649
Women aged 35-60 yr
old
Blood
Blood was measured by
AAS
Age at Measurement:
35-60 yr
Geometric mean:
1.4 |jg/dL Mean:
1.6 Mg/dL
Max: 17.0 pg/dL
Female reproductive
function: Serum FSH and
LH
Serum FSH and LH were
measured using a
microparticle enzyme
immunoassay
Age at outcome:
35-60 yr
Regression analyses:
the slopes were
adjusted forage, Iog10
serum bone alkaline
phosphatase, log 10
urine N-telopeptides,
log 10 serum cotinine,
alcohol use, currently
breastfeeding,
hysterectomy, one
ovary removed, Depo-
Provera use, medical
conditions or
treatments, hormone
pill use, and hormone
patch use
(3 (95% Cl)b, as slope for
serum FSH and LH per
log 10 blood Pb increase
Serum FSH (IU/L)
Post-menopausal: 26.38
(13.39, 39.38)
Pregnant: -0.08 (-1.11,
0.95)
Menstruating: 1.50 (-2.29,
5.30)
Both ovaries removed:
27.71 (1.64, 53.78)
Birth control pills: -0.33
(-6.52, 5.86)
Pre-menopausal: 11.97
(3.27, 20.66)
Serum LH (IU/L)
Post-menopausal: 11.63
(4.40, 18.86)
Pregnant: 2.12 (-14.62,
18.86)
Menstruating: 0.87 (-2.20,
3.94)
Both ovaries removed:
20.59 (2.14, 39.04)
Birth control pills: 2.19
(-1.35, 5.72)
Pre-menopausal: 7.44
(-0.26, 15.14)
8-210
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Chen etal. (2016)
Shanghai, Jiangxi
Province and
Zhejiang Province
China
2014
Cross-sectional
SPECT-China
n: 2286 men and 1571
postmenopausal women
SPECT-China is a
population-based cross-
sectional survey on the
prevalence of metabolic
diseases and risk factors
in East China. Men and
postmenopausal women
(age >55 yr) who were
not taking hormone
replacement therapy,
without a history of
hysterectomy and
oophorectomy were
recruited.
Blood
Blood was measured by
AAS
Age at measurement:
Median age 63 (IQR: 59-
68)
Median0: 4.1 |jg/dL
75thc: 5.981 pg/dL
Quartile0 (pg/dL)
Q1
Q2
Q3
Q4
<2.7
2.7-4.099
4.1-5.980
>5.980
Female reproductive
function: Reproductive
hormone levels
Venous blood samples were
drawn from all subjects after
an overnight fast of at least
8 hr. HbA1c was assessed
via high-performance liquid
chromatography (MQ-
2000PT, China). tT, E2, LH
and FSH levels were
measured using
chemiluminescence assays
(Siemens Immulite 2000,
Germany). SHBG levels
were detected using Cobas
e601
electrochemiluminescence
immunoassays (Roche,
Switzerland).
Age at outcome:
Median age 63 (IQR: 59-68)
Linear regression
models were adjusted
for age, current
smoking status, BMI,
SBP, diabetes, and
blood Cd level
(3 (SE)d
SHBG
Q1
Q2
Q3
Q4
tT
Q1
Q2
Q3
Q4
E2
Q1
Q2
Q3
Q4
Reference
0.010 (0.015)
0.018 (0.015)
0.048 (0.016)
Reference
-0.033 (0.019)
-0.017 (0.019)
-0.016 (0.020)
Reference
-0.001 (0.019)
-0.020 (0.019)
-0.021 (0.020)
FSH
Q1
Q2
Q3
Q4
Reference
0.013 (0.015)
0.047 (0.015)
0.046 (0.016)
LH
Q1
Q2
Q3
Q4
Reference
0.022 (0.015)
0.027 (0.016)
0.037 (0.016)
8-211
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Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Lee etal. (2019)
Busan
Korea
2012-2014
Cross-sectional
Second Korean National Blood
Environmental Health
Survey
n: 4,689 adults
2,763 men and 1,926
postmenopausal women
aged 50 yr or over
Blood was measured by
GFAAS
Age at Measurement:
50 yr or older
Median: 2.05 |jg/dL
75th: 2.67 pg/dL
Female reproductive
function: Follicle-stimulating
hormone levels
Serum FSH levels were
measured using a
chemiluminescence
immunoassay
(chemiluminescent
immunoassay; ADVIA
Centaur XP; Siemens,
Tarrytown, NY, United
States)
Multiple linear
regression adjusted for
age, BMI, smoking
status, and alcohol
consumption
(3 (95% Cl)b: 2.929 (0.480,
5.377)
Mendola etal. (2013) NHANES
United States
1999-2010
Cross-sectional
n: 3,221 (2,158
menstruating and 1,063
menopause)
Women aged 45-55 yr
Blood
Blood was measured by
AAS in 1999-2002 and
ICP-MS in 2003-2010
Age at measurement: 45-
55 yr
Geometric mean:
Menopausal women:
1.71 pg/dL
Menstruating women
1.23 pg/dL
Quartiles (pg/dL)
Q1
Q2
Q3
Q4
LOD-1.0
1.0-1.4
1.4-2.1
2.1-22.4
Female reproductive:
Menopause
Menopause was
dichotomized: women with
at least one menstrual cycle
in the past 12 mo were
categorized as "No" and
those with natural
menopause were "Yes"
Age at outcome:
45-55 yr
Logistic regression
models were adjusted
for age, race/ethnicity,
current hormone use,
poverty, and smoking;
NHANES 1999-2002
models also adjusted
for bone alkaline
phosphatase; and
NHANES 2005-2008
models also adjusted
for femoral neck bone
density
OR (95% CI)
NHANES 1999-2010
Q1
Q2
Q3
Q4
Reference
1.7 (1.0, 2.8)
2.1 (1.2, 3.6)
4.3 (2.6, 7.2)
NHANES 1999-2002
Q1: Reference
Q2: 1.0 (0.3, 3.5)
Q3: 1.3 (0.4, 4.5)
Q4: 5.1 (1.4, 18.0)
Adjusted for bone alkaline
phosphatase
Q1
Q2
Q3
Q4
Reference
1.1 (0.3, 3.9)
1.2 (0.3, 4.7)
4.2 (1.2, 15.5)
NHANES 2005-2008
Q1: Reference
8-212
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Q2: 3.0 (0.9, 9.8)
Q3: 4.9 (1.5, 16.1)
Q4: 10.5 (3.1, 35)
Adjusted for femoral neck
bone density
Q1
Q2
Q3
Q4
Reference
3.4 (0.9, 12.2)
4.1 (1.1, 15.2)
9.7 (2.8, 33)
Eum etal. (2014)
Boston, MA
United States
1990-1994 (2001-
2004)
Cohort
Nurse's Health Study
n: 434
Female registered
nurses, 30 to 55 yr of age
and living in 11
U.S. states, completed a
questionnaire on their
medical history and
health-related behaviors;
analysis restricted to
women in the Boston
area who did not have a
history of a major,
chronic disease; and
were not obese from
1990-1994 and women
no history of chronic
diseases (no reported
diagnosis of
hypertension,
cardiovascular disease,
renal disease, diabetes,
or malignancies) invited
to participate from 2001
through 2004
Blood and bone
Bone was measured by
K-XRF at each woman's
mid-tibial shaft and
patella. Blood was
measured by GFAAS with
Zeeman background
correction
Age at measurement:
46 yr or older at the time
of bone Pb measurement
Median
Tibia: 10 |jg/g
Patella: 12 |jg/g
Blood: 3 |jg/dL
75th
Tibia: 15 |jg/g
Patella: 18 |jg/g
Blood: 4 |jg/dL
Tertiles
Tibia (|jg/g)
Female reproductive
function: Early menopause
Menopausal status was
determined on the first
Nurse's Health Study
questionnaire in 1976 and
then again on each biennial
questionnaire by asking
whether the participants'
menstrual periods had
ceased permanently; early
menopause as natural
menopause occurring
before 45 yr of age
Age at outcome:
Age at reporting of
menopausal status
Ordinary least-squares
linear regression to
analyze age at
menopause adjusted
for sub-study group,
age at bone Pb
measure, age at bone
Pb measure squared,
year of birth, age at
menarche, months of
oral contraceptive use,
parity, and pack-years
of smoking; logistic
regression for early
menopause adjusted
for sub-study group,
age at bone Pb
measure, age at bone
Pb measure squared,
year of birth, age at
menarche, months of
oral contraceptive use,
parity, and pack-years
of smoking
(3 (95% CI), as difference in
age at natural menopause
(year)
Tibia
T1
T2
T3
Reference
-0.80 (-1.67, 0.06)
-1.21 (-2.08, -.035)
p for trend: 0.006
Patella
T1
T2
T3
Reference
-0.32 (-1.18, 0.55)
-0.00 (-0.88, 0.87)
p for trend: 0.99
Blood
T1
T2
T3
Reference
0.08 (-0.80, 0.96)
-0.28 (-1.13, 0.56)
p for trend: 0.54
8-213
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
T1
<6.5
T2
6.513
T3
>13
Patella (pg/g)
T1
<8
T2
8-15
T3
>15
Blood (|jg/dL)
T1
<3
T2
3
T3
>3
Effects on Female Fertility
Lee et al. (2020)
United States
2013-2014 and
2015-2016
Cross-sectional
NHANES (2013-2014
and 2015-2016)
n: 124
Women aged 20-39 yr
without a history of
hysterectomy and/or
bilateral oophorectomy
Blood
Blood was measured by
ICP-MS
Age at Measurement:
20-39 yr
Geometric mean:
0.50 |jg/dL (95% CI: 0.43,
0.57)
Female reproductive
function: Female infertility
Infertility is defined as the
absence of pregnancy with
unprotected intercourse for
1 yr and was assessed
through a self-reported
questionnaire
Age at outcome:
20-39 yr
Logistic regression
analyses were adjusted
for age, ethnicity,
annual family income,
education, marital
status, smoking history,
alcohol consumption,
physical activity, and
BMI
OR (95% Cl)b: 2.60 (1.05,
6.41) per 2-fold increase in
BLLs
OR (95% CI)
T1
T2
T3
Reference
5.40 (1.47, 19.78)
5.62 (1.13, 27.90)
Tertiles (pg/dL)
T1
T2
T3
0.11-0.38
0.41-0.62
0.63-5.37
Louis etal. (2012) LIFE Studv
Blood
Female reproductive
Cox models for
OR (95% CI), as
n: 501
function: Fecundity
discrete survival time,
fecundability OR
Michigan (4 counties)
and Texas (12 Female ages 18-44 yr
Blood was measured by
ICP-MS
Women were instructed in
which is a proportional
odds model, adjusted
Female only exposure:
0.97 (0.85, 1.11)
8-214
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
counties)
United States
2005-2009
Cohort
and male ages >18 yr; in
a committed relationship;
ability to communicate in
English or Spanish;
menstrual cycles
between 21 and 42 d; no
hormonal contraception
injections during past
year; and no sterilization
procedures or physician
diagnosed infertility
Age at Measurement:
19-40 yr
Geometric mean
Pregnant female:
0.66 |jg/dL
Not pregnant female:
0.76 |jg/dL
Tertiles (pg/dL)
T1: 0.23-0.57
T2
T3
0.58-0.78
0.79-5.84
the use of the Clearblue
Easy fertility monitors
consistent with the
manufacturer's guidance
commencing on day six for
tracking daily levels of E3G
and LH. Women also used
the digital Clearblue Easy
home pregnancy test upon
enrollment to ensure the
absence of pregnancy at
study start and on the day
menses was expected for
each cycle under
observation in the study.
Age at outcome:
Average age with
pregnancy: 29.8
Average age without
pregnancy: 30.6
for age, BMI, cotinine,
parity, serum lipids,
and site
(Texas/Michigan)
Couple exposure:
Female exposure: 1.06
(0.91, 1.24)
Male exposure: 0.82 (0.68,
0.97)
Lai etal. (2017)
Taipei
Taiwan
2008-2010
Cross-sectional
n: 190 infertile women
including 68 patients with
endometriosis and 122
controls
Women who visited the
infertility clinic first time
for a specific
gynecologist at Taipei
Medical University
Hospital; women with
diagnoses such as
ovarian cyst, premature
ovarian failure, repeated
implantation failure or
pregnancy were
excluded
Blood
Blood was measured by
ICP-MS
Age at measurement:
Mean age for women with
endometriosis: 35.3 (SD:
4.1)
Mean age for women
without endometriosis:
35.3 (SD: 5.0)
Geometric mean0
Female reproductive
function: Endometriosis
among infertile women
Endometriosis status was
determined by laparoscopy
Age at outcome:
Mean age for women with
endometriosis: 35.3 (SD:
4.1)
Mean age for women
without endometriosis: 35.3
(SD: 5.0)
Multivariate logistic
regression adjusted for
age, body fat
proportion, educational
level, age at menarche,
and regularity of
menstrual cycle
OR (95% CI)
T1: Reference
T2: 1.73 (0.77,
T3: 2.59 (1.11,
3.88)
6.06)
8-215
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Outcome
Confounders
Effect Estimates and 95%
Clsa
Reference and
Study Design
Study Population
Exposure Assessment
Women with
endometriosis:
1.337 |jg/dL
Women without
endometriosis:
0.853 |jg/dL
Median0
Women with
endometriosis:
2.130 |jg/dL
Women without
endometriosis:
0.464 |jg/dL
Tertiles0 (|jg/dL)
T1: <0.38
T2: 0.38-3.05
T3: >3.05
Li et al. (2022)
Hefei
China
October 2019 -
January 2020
Cohort
n: 1184
Participants selected
from First Affiliated
Hospital of Anhui Medical
University while seeking
IVF treatment and
diagnosed infertility with
their partner. Inclusion
criteria: women were
aged between 20 and
45 yr; couples were
diagnosed with infertility
(failure to establish a
clinical pregnancy with
unprotected intercourse
for at least 1 yr); and IVF
indicators were tubal
Blood
Maternal blood (serum)
was measured by ICP-MS
Age at measurement:
Maternal age at collection
(day before oocytes were
retrieved for IVF); female
partner mean age was
30.22 yr
Geometric meane:
0.0877 |jg/dL
Mediane: 0.0924 |jg/dL
75the: 0.14399 pg/dL
Female reproductive
function - Effects on female
fertility: Fertility - successful
implantation, clinical
pregnancy
A serum hCG level
>25 mlU/mL on the 14th d
after embryo transfer was
considered as successful
implantation. Clinical
pregnancy was defined as
an ultrasound-confirmed
intrauterine pregnancy on
the 30th d after embryo
transfer.
Age at outcome:
Logistic regression
model for successful
implantation adjusted
for maternal age, BMI,
treatment protocol,
FSH levels, sperm
viability, cycle type,
and embryo quality.
Logistic regression
model for clinical
pregnancy adjusted for
maternal age, BMI,
treatment protocol,
endometrial thickness
on hCG day, and
embryo quality. Linear
regression models for
Mil rate, fertility rate,
OR (95%CI)b:
Successful implantation
Continuous: 0.85 (0.77,
0.94)
Tertiles
Low: Reference
Medium: 1.11 (0.75, 1.63)
High: 0.58 (0.40, 0.85)
Clinical pregnancy
Continuous: 0.95 (0.91,
0.99)
Tertiles
Low: Reference
Medium: 0.72 (0.37, 1.38)
8-216
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
factor, ovulation failure,
or other factors for
female partner or male
factor or unexplained
fertility.
Tertilese (pg/dL)
Low: 0.002-0.065
Medium: 0.065-0.125
High: 0.125-0.481
Female partner mean age:
30.22 yr
2PN rate, blastocyte
rate, and high-quality
embryo rate were
adjusted for maternal
age, BMI, education
level, infertility type,
FSH and sperm
concentration
High: 0.56 (0.29, 1.06)
(3 (95% Cl)b:
Mil rate: 0.090 (-0.024,
0.204)
Fertility rate: -0.033
(-0.151, 0.086)
2PN rate: -0.019 (-0.100,
0.062)
Blastocyst rate: 0.046
(-0.052, 0.144)
High quality embryo rate:
-0.143 (-0.322, -0.037)
Zhou et al. (2021a) n: 195
China
2018-2019
Cohort
Couples undergoing IVF.
Women with
endometriosis,
hydrosalpinx, abnormal
uterine cavity and men
with azoospermia, severe
oligozoospermia,
asthenospermia and
dysspermia were
excluded from the study.
Blood
Maternal blood (serum),
follicular fluid, and
seminal plasma from male
partner
Age at Measurement:
Female partner mean
age: 30.27 yr
Male partner mean age:
31.57 yr
Mean0
Maternal serum:
0.301 |jg/dL
Follicular fluid:
0.742 |jg/dL
Seminal plasma:
0.882 |jg/dL
Median0
Female reproductive
function - Effects on female
fertility: IVF outcome
The IVF outcomes included
were normal fertilization,
good embryo, blastocyst
formation, high-quality
blastocyst, pregnancy, and
live birth
Age at outcome:
Female partner mean age:
30.27 yr
Male partner mean age:
31.57 yr
Poisson regression
models were adjusted
for age and BMI
RR (95% Cl)b
Normal fertilization
Maternal serum: 0.94
(0.42, 1.93)
Follicular fluid: 0.82 (0.18,
2.39)
Seminal plasma: 1.55
(0.64, 3.3)
Good embryo
Maternal serum: 1.00
(0.36, 2.38)
Follicular fluid: 0.78 (0.09,
3.03)
Seminal plasma: 1.86
(1.05, 3.11)
Blastocyst formation
Maternal serum: 1.06 (0.2,
3.91)
Follicular fluid: 0.41 (0,
3.63)
8-217
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Outcome
Confounders
Effect Estimates and 95%
Clsa
Seminal plasma: 1.77
(0.78, 3.58)
High-quality blastocyst
Maternal serum: 1.68
(0.15, 9.43)
Follicular fluid: 0.35 (0,
7.11)
Seminal plasma: 2.66
(0.67, 8)
Pregnancy
Maternal serum: 0.18
(0.01, 1.91)
Follicular fluid: 0.01 (0,
0.03)
Seminal plasma: 0.04 (0,
1.45)
Reference and
Study Design
Study Population
Exposure Assessment
Maternal serum:
0.245 |jg/dL
Follicular fluid:
0.178 |jg/dL
Seminal plasma:
0.486 |jg/dL
75thc
Maternal serum:
0.317 |jg/dL
Follicular fluid:
0.326 |jg/dL
Seminal plasma:
1.245 |jg/dL
Live birth
Maternal serum: 0.25
(0.01, 2.8)
Follicular fluid: 0 (0, 0.09)
Seminal plasma: 0.01 (0,
1.08)
Effects on Morphology or Histology of Female Sex Organs (Ovaries, Uterus, Fallopian Tubes/Oviducts, Cervix, Vagina, and/or Mammary Glands)
Ye et al. (2017)
Seoul
South Korea
September to
November 2014
n: 288 (46 with fibroids
and 242 without)
Premenopausal women
between 30 and 49 yr
old, who were not
pregnant or
breastfeeding, whose
heavy metal levels at the
Blood
Blood was measured by
GFAAS
Age at Measurement:
30-49 yr
Female reproductive
function - Effects on
morphology and histology of
female sex organs: Uterine
fibroids
Diagnosis of uterine fibroids
was based on pelvic
ultrasonography and two
Logistic regression
models adjusted for
age, BMI, gravidity,
oral contraceptive pill
administration history,
regularity of menstrual
cycle, hemoglobin
level, and serum
cotinine levels; linear
OR (95% CI)b
Presence of uterine
fibroids: 1.39 (0.75, 2.56)
(3 (95% Cl)b
Volume of uterine fibroids:
0.12 (-2.26, 2.51)
8-218
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Cross-sectional
time might have been
influenced by these
circumstances and might
have been less
representative of heavy
metal levels at the time of
diagnosis, and who had
received hysterectomies
Geometric mean:
1.36 [jg/dL
Quartiles (pg/dL)
Q1: <1.1
Q2: 1.1-1.3
Q3: 1.3-1.8
Q4: 1.8-3.2
questions
Age at outcome:
30-49 yr
regression models
were adjusted for age,
BMI, gravidity, oral
contraceptive pill
administration history,
regularity of menstrual
cycle, hemoglobin
level, and serum
cotinine levels
Q1
Q2
Q3
Q4
Reference
-0.42 (-2.69, 1.85)
0.85 (-1.67, 3.37)
-1.23 (-3.74, 1.29)
2PN = oocytes with two pronuclei; AAS = atomic absorption spectrometry; BMI = body mass index; d = day(s); E2 = estradiol; E3G = estrone-3-glucuronide; FSH = follicle stimulating
hormone; GFAAS = graphite furnace atomic absorption spectrometry; hCG = human chorionic gonadotropin; ICP-MS = inductively coupled plasma mass spectrometry;
IQR = interquartile range; IVF = in vitro fertilization; K-XRF = K-shell X-ray fluorescence instrument; LH = luteinizing hormone; LIFE = Longitudinal Investigation of Fertility and the
Environment; LOD = limit of detection; Mil = metaphase II; mo = month(s); NHANES = National Health and Nutrition Examination Survey; OR = odds ratio; SBP = systolic blood
pressure; SD = standard deviation; SHBG = sex hormone binding globulin; SPECT = Survey on the Prevalence in East China for Metabolic Diseases and Risk Factors; tT = total
testosterone; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bEffects estimates unable to be standardized.
°Pb measurements were converted from |jg/L to |jg/dL.
dNo CIs provided.
ePb measurements were converted from ng/Lto |jg/dL.
8-219
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Table 8-15
Animal toxicological studies of Pb exposure and female reproductive effects
Study
Species (Stock/Strain), n, Sex Timing of Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
Corv-Slechta et al.
(2013)
Mouse (C57BL/6) GD -61 to PND 365
Control (untreated), F, n = 16
100 ppm, F, n = 16
Dams were dosed starting 2 mo
prior to mating. Offspring were
continued on the same
exposure as their dams until the
end of the experiment at 12 mo
of age.
0.22 |jg/dL for control
dams at weaning
12.12 |jg/dL for
100 ppm dams at
weaning
Litter Size,
Maternal
Body Weight
Weston et al. (2014) Rat (Lona-Evansl GD -76 to PND 21
Dams
Control (untreated), F, n = 20
50 ppm Pb, F, n = 19
Dams were dosed via drinking
water starting 2-3 mo prior to
breeding. Exposure ended at
weaning (PND 21).
Dams (PND 21):
0.500 |jg/dL for control
7.72 |jg/dL for 50 ppm
Pb
Litter Size,
Number of
Litters
Pups
Control (untreated), M/F, n = 12.4
(7/5.4 average number of male and
female pups per litter in control)
50 ppm Pb, M/F, n = 7.4 (6.3/1.1
average number of male and
female pups per litter in Pb NS
group)
Pups (PND 5-6):
0.603 |jg/dL for control
males
0.690 |jg/dL for control
females
15.7 |jg/dL for 50 ppm
Pb males
14.6 |jg/dL for 50 ppm
Pb females
8-220
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Study
Species (Stock/Strain), n, Sex Timing of Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
Betharia and Maher
(2012)
Rat (Sprague-Dawley)
Control (untreated), F, n = 6 dams
10 |jg/mL Pb, F, n = 6 dams
GD Oto PND 20
Dams were dosed via drinking
water throughout pregnancy
until weaning (PND 20).
Pups:
PND 2
0.188 |jg/dL for control
9.03 |jg/dL for 10 |jg/ml_
Pb
Litter Size
PND 25:
0.088 |jg/dL for control
0.976 |jg/dL for
10 |jg/ml_ Pb
PND 60:
0.0244 |jg/dL for control
0.0318 |jg/dL for
10 |jg/ml_ Pb
Schneider et al. (2016) Mouse (C57BL/6)
Control (untreated), F, n = NR
100 ppm Pb, F, n = NR
GD -61 to PND 21
Dams were dosed via drinking
water starting 2 mo prior to
mating through lactation
(weaning assumed to be
PND 21).
Dams were also treated to a
non-stress or prenatal stress
condition. Only data from dams
in the non-stress condition were
used.
Dams at weaning
(assumed PND 21):
0.22 |jg/dL for control
12.61 |jg/dL for
100 ppm Pb
Pups (PND 5-6):
0.37 |jg/dL for control
10.2 |jg/dL for 100 ppm
Pb
Maternal
Body
Weight,
Litter Size
Saleh et al. (2018)
Rat (Sprague-Dawley)
Control (vehicle), F, n =
160 ppm Pb, F, n = 8
320 ppm Pb, F, n = 8
GD 1 to 20
Dams were dosed via oral
gavage. Authors report a
significant decrease in brain
weight occurred, indicating
potential overt toxicity.
Dams (GD 20):
5.1 |jg/dL for control
27.7 |jg/dL for 160 ppm
Pb
41.5 |jg/dL for 320 ppm
Pb
Maternal
Body Weight
8-221
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Study
Species (Stock/Strain), n, Sex
Timing of Exposure
Exposure Details
(Concentration, Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
Baranowska-Bosiacka
Rat (Wistar)
GD 1 to PND 21
Dams were exposed via
NR for Dams
Sex Ratio
etal. (2013)
Control (untreated), F, n = 3 dams
drinking water throughout
0.1% Pb, F, n = 3 dams
pregnancy until weaning
(PND 21).
Pups (PND 28):
Control, M/F, n = 36 (17/19) pups
0.93 |jg/dL for control
0.1% Pb, M/F, n = 36 (18/18) pups
6.86 |jg/dL for 0.1% Pb
Saleh et al. (2019)
Rat (Sprague-Dawley)
Control (vehicle), F, n = 8 dams
160 ppm Pb, F, n = 8 dams
320 ppm Pb, F, n = 8 dams
GD 1 to 20
Dams were dosed via oral
gavage. Authors report a
significant decrease in brain
weight occurred, indicating
potential overt toxicity.
Dams (GD 20):
5.26 |jg/dL for control
23.9 |jg/dL for 160 ppm
Pb
42.9 |jg/dL for 320 ppm
Pb
Maternal
Body Weight
BLL = blood lead level; F = female; GD = gestational day; M = male; mo = month(s); NR = not reported; Pb = lead; PND = postnatal day; NS = non-stress.
8-222
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Table 8-16
Epidemiologic studies on exposure to Pb and male reproductive effects
Reference and
Study Design
Study Population Exposure Assessment Outcome Confounders
Effect Estimates and 95%
Clsa
Effects on Sperm/Semen Production, Quality, and Function
Li etal. (2015)
Taiwan
May 2012 to
February 2013
Cross-sectional
n: 154
Male participants were
recruited from a
reproductive medical
center and did not have
obstructive azoospermia,
cryptorchidism,
varicoceles, hydrocele,
orchitis, or epididymitis;
did not have testicular
injury or underwent
testicular surgery before
the study period
Blood
Blood was measured by
ICP-MS
Age at Measurement:
Mean age: 34.8 yr
Mean (SD)b: 2.78 (1.85)
pg/dL
Male reproductive effects:
Seminal parameters
From semen samples the
following parameters were
assessed: sperm
concentration, semen
volume, number of sperm,
percentage of total motility
sperm, percentage of
progressive motility sperm,
and percentage of sperm
with normal morphology
Age at outcome:
Mean age: 34.8 yr
Multiple logistic
regression models were
adjusted for FSH, LH,
prolactin, and
testosterone were input
into the model and then
adjusted for age and
education
OR (95% CI)
Low-quality semen: 1.040
(1.011, 1.069)
Sperm concentration: 1.046
(1.015, 1.078)
Numbers of sperm: 1.041
(1.012, 1.071)
Total motility sperm: 1.057
(1.026, 1.089)
Progressive motility sperm:
1.047 (1.014, 1.080)
Sperm with normal
morphology: 1.071 (1.025,
1.118)
Sukhn etal. (2018)
Beirut
Lebanon
January 2003 and
December 2009
Cross-sectional
Environment and Male
Infertility study
n: 116
Male partners of infertile
heterosexual couples who
attended the fertility clinic
at the American University
of Beirut Medical Center
were recruited. Men were
18 to 55 yr of age, had a
BMI of 18 to 30 kg/m2, and
had not been on any
hormone therapy for the
past 6 mo, no diabetes,
endocrine disease,
fertility-related genetic
disorders, obstructive
azoospermia,
Blood and other: seminal
fluid
Blood and seminal fluid
were measured by ICP-
MS equipped with a cell
dynamic range
Age at Measurement:
18-55yr
Meanb
Blood
Overall: 3.121 pg/dL
Low-quality semen group:
5.198 |jg/dL
Male reproductive effects:
Semen quality
Participants with a semen
volume <1.5 mL, sperm
concentration
<15 million/mL, total count
<39 million, progressive
motility <32%, viability
<58%, and/or normal
WHO morphology <30%
were assigned to the low
quality semen group A.
Participants whose semen
analyses expressed better
results in all the above
parameters were assigned
to the normal-quality
semen group B. Sperm
Logistic regression;
age, cigarette smoking,
alcohol intake, and
period of sexual
abstinence
OR (95% CI)
Blood
Volume (<1.5 mL)
Q1
Q2
Q3
Q4
Reference
0.53 (0.11, 2.44)
0.24 (0.02, 2.24)
1.32 (0.33, 2.56)
p for trend: 0.26
Concentration (<15 M/mL)
Q1
Q2
Q3
Q4
Reference
0.51 (0.16, 1.63)
1.17 (0.37, 3.73)
1.58 (0.53, 4.68)
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Confounders
Effect Estimates and 95%
Clsa
cryptorchidism, varicocele, Normal quality semen
hydrocele, orchitis.
Epididymitis, and/or
history of testicular injury
or surgery
group 3.575 [jg/dL
Seminal fluid
Overall: 0.540 [jg/dL
Low-quality semen group:
1.626 [jg/dL
Normal quality semen
group: 1.285 |jg/dL
Medianb
Blood
Low-quality semen group:
3.257 [jg/dL
Normal quality semen
group: 3.098 [jg/dL
Seminal fluid
Low-quality semen group:
0.588 [jg/dL
Normal quality semen
group: 0.470 |jg/dL
Quartilesb (pg/dL)
concentration (million/mL)
and progressive motility
(%) were determined
manually using a Makler®
counting chamber. Total
sperm count (million) was
calculated as sperm
concentration * semen
volume. Sperm
morphology was
determined by high-power
magnification (* 1000) on
air-dried smears stained
with a Wright Giemsa stain
based on the WHO
guidelines.
Age at outcome:
18-55 yr
Q1
Q2
Q3
Q4
LOD-2.199
2.200-3.256
3.257-5.357
>5.358
p for trend: 0.26
Total count (<39 M)
Q1
Q2
Q3
Q4
Reference
0.36 (0.11, 1.18)
0.83 (0.26, 2.65)
1.35 (0.46, 3.96)
p for trend: 0.15
Progressive motility (<32%)
Q1
Q2
Q3
Q4
Reference
0.70 (0.19, 2.62)
0.78 (0.19, 3.19)
1.47 (0.43, 5.02)
p for trend: 0.66
Viability (<58%)
Q1
Q2
Q3
Q4
Reference
0.44 (0.14, 1.39)
0.68 (0.21, 2.21)
1.35 (0.46, 3.96)
p for trend: 0.23
WHO morphology (<30%)
Q1
Q2
Q3
Q4
Reference
0.50 (0.15, 1.66)
0.93 (0.28, 3.10)
0.84 (0.26, 2.66)
p for trend: 0.68
Seminal Fluid
Blood
Volume (<1.5 mL)
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Confounders
Effect Estimates and 95%
Clsa
Q1
Q2
Q3
Q4
Reference
0.86 (0.16, 4.67)
1.34 (0.25, 7.17)
2.07 (0.37, 11.51)
p for trend: 0.95
Concentration (<15 M/mL)
Q1
Q2
Q3
Q4
Reference
1.57 (0.50, 4.92)
1.99 (0.62, 6.38)
1.94 (0.59, 6.35)
p for trend: 0.64
Total count (<39 M)
Q1
Q2
Q3
Q4
Reference
1.66 (0.51, 5.46)
3.33 (1.01, 10.99)
2.00 (0.58, 6.85)
p for trend: 0.24
Progressive motility (<32%)
Q1
Q2
Q3
Q4
Reference
4.36 (0.83, 22.81)
6.35 (1.21, 33.19)
2.40 (0.39, 14.49)
p for trend: 0.09
Viability (<58%)
Q1
Q2
Q3
Q4
Reference
8.00 (1.59, 40.30)
12.00 (2.34, 61.52)
10.15 (1.95, 52.92)
p for trend: 0.006
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Confounders
Effect Estimates and 95%
Clsa
WHO morphology (<30%)
Q1
Q2
Q3
Q4
Reference
3.83 (0.924, 15.90)
6.57 (1.57, 27.43)
2.02 (0.426, 9.55)
p for trend: 0.06
Shi etal. (2021)
Hong Kong
November 2015-
November 2016
Cross-sectional
n: 288
Male subjects who
underwent SA as part of
the fertility assessment at
the andrologyjaboratory
of Prince of Wales
Hospital. Participants were
excluded with medical
conditions azoospermia;
andrological conditions
(which are known to affect
semen parameters
including genetic
conditions); history of
mumps orchitis, severe
varicocele, undescended
testis; history of testicular
torsion or scrotal injury,
congenital bilateral
absence of vas deferent,
and urogenital infections;
taking medication known
to affect semen
parameters, including
steroid, finasteride,
calcium channel blockers;
history of malignant
disease; known mental
disorders; drug abuse;
failure to complete the
Blood
Blood was measured by
ICP-MS
Age at Measurement
Mean age: 37.9 yr
Geometric meanb:
3.175 |jg/dL
Medianb: 2.719 |jg/dL
75thb: 3.437 pg/dL
Quartilesb (pg/dL)
Q1
Q2
Q3
Q4
<2.159
>2.159-2.719
>2.719-3.437
>3.437
Male reproductive effects:
Seminal parameters
Semen volume was
measured by a wide-bore
graduated pipette_with the
graduation of 0.1-
ml. Sperm
concentration and motility
were examined under
a phase contrast
microscope with the
magnification of * 200 or
400. Diff-Quik staining kit
(Dade Behring AG,
Dudingen, Switzerland)
and Tygerberg Strict
Criteria were used to
evaluate the sperm
morphology. Sperm DNA
fragmentation was
measured by sperm
chromatin structure assay.
Age at outcome:
Mean age: 37.9 yr
Multivariable linear
regression adjusted for
(1) male age and daily
coffee intake for semen
volume models; (2)
abstinence time,
average sleep duration
for sperm concentration
models; (3) male age,
abstinence time, and
daily coffee intake for
total sperm count
models; (4) male age
and daily juice intake
for the sperm motility
models; (5) male age
and abstinence time for
total motility count
models; (6) no
adjustment for normal
morphology or sperm
vitality models; (7) male
age, abstinence time,
and irregular sleeping
habit for DNA
fragmentation index
models; (8) daily juice
intake for percentage of
acrosome reacted
sperm models.
B (95% CI)
Semen volume
Q1
Q2
Q3
Q4
Reference
-0.05 (-0.70, 0.37)
0.04 (-0.39, 0.65)
0.08 (-0.32, 0.83)
p for trend: 0.48
Sperm concentration
Q1
Q2
Q3
Q4
Reference
0.02 (-0.45, 0.58)
-0.02 (-0.57, 0.43)
-0.10 (-0.85, 0.26)
p for trend: 0.34
Total sperm count
Q1
Q2
Q3
Q4
Reference
-0.01 (-0.60, 0.53)
-0.03 (-0.66, 0.43)
-0.05 (-0.76, 0.44)
p for trend: 0.55
Sperm motility
Q1: Reference
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Effect Estimates and 95%
Clsa
lifestyle questionnaire; and
refusal to donate blood or
semen samples.
Q2
Q3
Q4
-0.09 (-10.21, 3.02)
-0.15 (-12.47, 0.49)
-0.08 (-10.26, 4.02)
p for trend: 0.77
Total motility count
Q1
Q2
Q3
Q4
Reference
-0.07 (-0.97, 0.38)
-0.08 (-0.97, 0.36)
-0.12 (-1.20, 0.25)
p for trend: 0.16
Normal morphology
Q1
Q2
Q3
Q4
Reference
-0.13 (-1.16, 0.13)
-0.20 (-1.43, -0.16)
-0.20 (-1.52, -0.10)
p for trend: 0.20
Sperm vitality
Q1
Q2
Q3
Q4
Reference
0.12 (-0.04, 0.17)
0.01 (-0.10, 0.12)
-0.13 (-0.19, 0.04)
p for trend: 0.13
Percentage of acrosome
reacted sperm
Q1
Q2
Q3
Q4
Reference
-0.22 (-18.60, 0.97)
-0.05 (-11.60, 7.71)
-0.12 (-15.70, 5.79)
p for trend: 0.75
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Confounders
Effect Estimates and 95%
Clsa
Pant etal. (2014) n: 60
New Delhi
India
Cross-sectional
Male partners of couples
age 21-40 yr old attending
the Andrology Laboratory
of the Reproductive
Biology Department, All
India Institute of Medical
Sciences, New Delhi,
India for semen analysis
to assess their inability to
achieve a pregnancy were
selected.
Other: Semen
Semen measured by ICP-
AES
Age at Measurement:
mean age: 31.81 (SD:
5.27)
Mean (SD): 6.18
(2.16) pg/dL
Male reproductive effects:
Semen quality
Semen of volunteers was
collected and analyzed the
protocols of the WHO.
Sperm morphology was
determined according to
Kruger's strict criteria.
Comet assay: prepared
sperm samples were
observed under a
fluorescence microscope
with a total of 100 cells
were scored. The
percentage of tail DNA, tail
length, and tail moment
was evaluated by the
CometScore software
image analysis system.
Age at outcome:
mean age: 31.81 (SD:
5.27)
Multiple regressions,
adjusted for toxicants
(Cd, diethyl phthalate,
dibutyl phthalate, di[2-
ethylhexyl] phthalate),
age, BMI, tobacco,
smoking, alcohol, and
diet
(3 (95% Cl)c
Sperm motility (%): 2.43
(-4.87, -0.001)
Sperm concentration
(106/ml): -1.97 (-3.16,
-0.33)
Tail length: 3.79 (0.56, 7.02)
Percent DNA in tail: 1.31
(0.172, 3.74)
Tail moment: 1.20 (0.23,
2.16)
Jia et al. (2022)
Henan Province
China
December 2017 to
August 2018
Cross-sectional
n: 841
Males ranging from 18 to
50 yr of age with no
history of testicular injury,
urologist diagnosed
inflammation of the
urogenital system; history
of epididymitis; treatment
history of varicocele;
history of incomplete
orchiocatabasis or any of
the following that was
detected by an urologist at
physical examination:
Other: Semen
Seminal plasma was
measured by analyzed
using the kinetic energy
discrimination-based
Thermo iCAP Q ICP-MS
Age at Measurement
Mean ± SD:
29.55 ±5.45 yr
Median: 1.70 ppb
Male reproductive effects:
Seminal parameters
Semen of volunteers was
collected and analyzed the
protocols of the WHO.
Computer-assisted sperm
analysis technology was
used to analyze the
collected semen samples.
The quality indicators were
complete liquefaction,
semen volume, sperm
concentration, total sperm
count, progressive motility,
Multilinear regression
models were adjusted
for age, BMI, smoking,
and alcohol
consumption
(3 (95% Cl)c, per increase in
In-Pb seminal plasma
Semen volume: -0.10
(-0.27, 0.07)
Sperm concentration: 1.83
(-4.45, 8.12)
Total sperm number: 0.80
(-17.61, 19.21)
Progressive motility: 0.06
(-2.09, 2.21)
Normal morphological rate:
-0.04 (-0.41, 0.34)
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Confounders
Effect Estimates and 95%
Clsa
absence of prominentia
laryngea, absence of
pubes, abnormal breast,
absence of testis,
abnormal penis,
epididymal knob, or
varicocele.
75th: 2.36 ppb
non-progressive motility,
sperm motility, and sperm
motility parameters, such
as curve line velocity
(|jm/s), straight line
velocity (|-im/s), velocity of
average path (|-im/s),
lateral head movement
(amplitude of lateral head
displacement, |jm),
average motion degree (°),
linearity (%), straightness
(%), wobble, and beat
cross frequency (beat
cross frequency, Hz).
Curve line velocity: 0.35
(-1.17, 1.88)
Straight line velocity: 0.54
(-0.50, 1.58)
Velocity of average path:
0.37 (-0.96, 1.70)
Linearity: 0.49 (-0.88, 1.86)
Straightness: 0.46 (-1.07,
1.99)
Wobble: 0.22 (-1.33, 1.77)
Average motion degree:
-0.33 (-1.22, 0.56)
Beat cross frequency: 0.01 (-
0.14, 0.15)
Lateral head movement:
-0.04 (-0.11, 0.03)
Williams etal. (2022)
Russia
2003-2005 (follow-
up annually for
10 yr)
Cohort
Russian Children's Study
n: 223
Boys enrolled at age 8-
9 yr in 2003-2005 and
followed them annually for
10 yr.
Blood
Blood was measured by
Zeeman background
corrected flameless
GFAAS
Age at measurement: 8-
9 yr
Median: 3 |jg/dL
75th: 5 |jg/dL
Categories
Lower: <5 |jg/dL
Higher: >5 |jg/dL
Tertiles
Male reproductive effects:
Semen parameters
All semen samples were
assessed by a single
andrology technician and
analyzed according to
criteria of the Nordic
Association for
Andrology and European
Society of Human
Reproduction and
Embryology-Special
Interest Group in
Andrology and serum
hormonal levels were
analyzed using the
Architect i1000SR and
chemiluminescent
Mixed effect linear
regression models
adjusted for boys' BW,
total caloric intake, HTZ
at entry, breastfeeding
duration, monthly
household income, and
abstinence time
(3 (95% Cl)b, as adjusted
mean
Semen volume (mL)
Continuous, per log-blood
Pb: -0.40 (-0.82, 0.03)
Categories
Lower: 2.83 (2.61, 3.06)
Higher: 2.60 (2.27, 2.93)
Tertiles
Low: 2.92 (2.50, 3.34)
Medium: 2.79 (2.52, 3.06)
High: 2.60 (2.27, 2.93)
p for trend: 0.24
Sperm concentration
(mill/mL)
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Outcome
Confounders
Effect Estimates and 95%
Clsa
Low: <2 [jg/dL
Medium: 3-4 [jg/dL
High: >5 [jg/dL
microparticle
immunoassay.
Age at outcome: 18 yr or
older
Continuous, per log-blood
Pb: 0.09 (-0.13, 0.31)
Categories
Lower: 47.0 (41.3, 53.4)
Higher: 49.0 (37.8, 63.4)
Tertiles
Low: 41.3 (33.2, 51.3)
Medium:50.3 (42.8, 59.0)
High: 49.1 (38.0, 63.6)
p for trend: 0.33
Total sperm count (mill)
Continuous, per log-blood
Pb: -0.02 (-0.27, 0.23)
Categories
Lower: 111 (95.6, 129)
Higher: 107 (80.0, 143)
Tertiles
Low: 99 (76.4, 128)
Medium: 118 (95.5, 141)
High: 107 (80.3, 143)
p for trend: 0.68
Progressive sperm motility
(%)
Continuous, per log-blood
Pb: 1.77 (-0.55, 4.08)
Categories
Lower: 53.2 (51.7, 54.7)
Higher: 53.1 (50.9, 55.2)
Tertiles
Low: 51.2 (48.6, 53.9)
Medium: 54.3 (52.5, 56.1)
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Confounders
Effect Estimates and 95%
Clsa
High: 53.2 (50.9, 55.3)
p for trend: 0.29
Effects of Hormone Levels
Total progressive motile
sperm count (mill)
Continuous, per log-blood
Pb: 0.01 (-0.27, 0.29)
Categories
Lower: 57.7 (48.9, 68.1)
Higher: 55.7 (40.6, 76.4)
Tertiles
Low: 49.4 (36.8, 66.2)
Medium: 62.6 (51.3, 76.5)
High: 56.0 (40.8, 76.8)
p for trend: 0.57
Low semen quality
(probability)
Continuous, per log-blood
Pb: 0.20 (-0.22, 0.65)
Categories
Lower: 0.51 (0.44, 0.58)
Higher: 0.49 (0.39, 0.59)
Tertiles
Low: 0.43 (0.31, 0.55)
Medium: 0.55 (0.46, 0.63)
High: 0.49 (0.39, 0.59)
p for trend: 0.43
Kresovich et al. NHANES Blood Male reproductive effects: Linear regression (3 (SE)d
(2015) n: 86g Hormones models were adjusted Testosterone (ng/mL)
for age, BMI, race, „ *
a Q1: Reference
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Confounders
Effect Estimates and 95%
Clsa
United States
1999-2004
Cross-sectional
Males who were aged
>20 yr, no reported steroid
or thyroid mediation use,
and no reported thyroid
disease.
Blood was measured by
AAS (1999-2002) or ICP-
MS (2003-2004).
Age at measurement:
>20 yr
Median (weighted):
2.0 [jg/dL
75th: 2.8 pg/dL
Quartiles (pg/dL)
Testosterone,
androstanedione
glucuronide, and SHBG
were measured in blood
serum, and E2 in plasma.
All sex hormones were
detected by immunoassay.
Age at outcome: >20 yr
diabetes status
(including prediabetes),
smoking status, and
alcohol intake; and Cd
Q1
Q2
Q3
Q4
<1.40
1.40-2.10
2.10-3.20
>3.20
Q2
Q3
Q4
0.39 (0.21)
0.56 (0.22)
0.81 (0.20)
p for trend: 0.0008
E2 (pg/mL)
Q1
Q2
Q3
Q4
Reference
-0.01 (0.03)
-0.01 (0.04)
-0.01 (0.04)
p for trend: 0.7849
IT (ng/dL)
Q1
Q2
Q3
Q4
Reference
0.83 (0.47)
0.55 (0.48)
0.81 (0.48)
p for trend: 0.2374
fE2 (pg/ml)
Q1
Q2
Q3
Q4
Reference
-0.01 (0.03)
-0.02 (0.04)
-0.03 (0.04)
p for trend: 0.4428
Androstanedione
glucuronide (ng/mL)
Q1: Reference
Q2: 0.03 (0.03)
Q3: -0.01 (0.03)
Q4: 0.02 (0.04)
p for trend: 0.8917
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Confounders
Effect Estimates and 95%
Clsa
SHBG (nmol/L)
Q1
Q2
Q3
Q4
Reference
0.01 (0.02)
0.05 (0.02)
0.05 (0.02)
p for trend: 0.0187
Adjusted for Cd
Testosterone (ng/mL)
Q1
Q2
Q3
Q4
Reference
0.38 (0.23)
0.54 (0.21)
0.79 (0.22)
p for trend: 0.0026
E2 (pg/mL)
Q1
Q2
Q3
Q4
Reference
0.00 (0.03)
0.01 (0.04)
0.02 (0.04)
p for trend: 0.6600
IT (ng/dL)
Q1
Q2
Q3
Q4
Reference
0.95 (0.50)
0.70 (0.51)
1.06 (0.51)
p for trend: 0.1388
fE2 (pg/ml)
Q1: Reference
Q2: 0.01 (0.03)
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Effect Estimates and 95%
Clsa
Q3: 0.00 (0.04)
Q4: 0.01 (0.04)
p for trend: 0.9456
Androstanedione
glucuronide (ng/mL)
Q1: Reference
Q2: 0.03 (0.03)
Q3: -0.02 (0.03)
Q4: 0.01 (0.04)
p for trend: 0.7620
SHBG (nmol/L)
Q1: Reference
Q2: -0.01 (0.02)
Q3: 0.03 (0.02)
Q4: 0.03 (0.03)
p for trend: 0.1333
Lewis and Meeker
(2015)
United States
2011-2012
Cross-sectional
NHANES
n: 484
Men that were 18-55 yr
old, that had complete
data on the metals of
interest, serum
testosterone, BMI, PIR,
race, serum cotinine, or
urinary creatinine
Blood
Blood was measured by
inductively coupled
dynamic reaction-plasma
mass spectrometry
Age at Measurement:
18-55yr
Geometric mean:
1.06 |jg/dL
75th: 1.59 pg/dL
Male reproductive effects:
Testosterone
Serum testosterone (total)
were measured by isotope
dilution-high performance
liquid chromatography-
tandem mass
spectrometry
Age at outcome:
18-55 yr
Multiple linear
regression, adjusted for
age, BMI, PIR, race,
and serum cotinine
(3 (95% Cl)c, as percent
change in serum
testosterone associated with
a doubling (100% increase)
in blood Pb concentration:
6.65 (2.09, 11.41)
Chen et al. (2016) SPECT-China Blood Male reproductive effects: Linear regression (3 (SE)d
n: 2286 men Reproductive hormone models were adjusted SHBG
Shanghai, Jiangxi levels for age and current
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Effect Estimates and 95%
Clsa
Province and
Zhejiang province
China
2014
Cross-sectional
SPECT-China is a
population-based cross-
sectional survey on the
prevalence of metabolic
diseases and risk factors
in East China. Men and
postmenopausal women
(age >55 yr) who were not
taking hormone
replacement therapy,
without a history of
hysterectomy and
oophorectomy were
recruited.
Blood was measured by
AAS
Age at Measurement:
Median (IQR) age: 54
(44-63)
Medianb: 4.400 |jg/dL
75thb: 6.230 pg/dL
Quartilesb (pg/dL)
Q1
Q2
Q3
<2.900
2.900-4.399
4.400-6.229
Q4: >6.229
Venous blood samples
were drawn from all
subjects after an overnight
fast of at least 8 hr. HbA1c
was assessed via high-
performance liquid
chromatography. tT, E2,
LH and FSH levels were
measured using
chemiluminescence
assays. SHBG levels were
detected using
electrochemiluminescence
immunoassays.
Age at outcome:
Median (IQR) age: 54 (44-
63)
smoking status, BMI,
SBP, diabetes and,
blood Cd level
Q1
Q2
Q3
Q4
Reference
<0.001 (0.011)
0.021 (0.011)
0.038 (0.012)
p for trend: <0.001
tT
Q1
Q2
Q3
Q4
Reference
0.001 (0.010)
0.010 (0.010)
0.033 (0.010)
p for trend: 0.001
E2
Q1
Q2
Q3
Q4
Reference
-0.008 (0.016)
0.014 (0.017)
-0.003 (0.017)
p for trend: 0.794
FSH
Q1: Reference
Q2: 0.010 (0.014)
Q3: 0.004 (0.014)
Q4: 0.030 (0.015)
p for trend: 0.067
LH
Q1: Reference
Q2: 0.018 (0.013)
Q3: 0.015 (0.013)
Q4: 0.028 (0.013)
p for trend: 0.065
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Outcome
Confounders
Effect Estimates and 95%
Clsa
Effects on Fertility
Louis etal. (2012)
Michigan (4
counties) and Texas
(12 counties)
United States
2005-2009
Cohort
LIFE Study
n: 501
Female ages 18-44 yr
and male ages >18 yr; in a
committed relationship;
ability to communicate in
English or Spanish;
menstrual cycles between
21 and 42 d; no hormonal
contraception injections
during past year; and no
sterilization procedures or
physician diagnosed
infertility
Blood
Blood was measured by
ICP-MS
Age at Measurement:
Average age for male
partner with pregnancy:
31.6 yr
Average age for male
partner without
pregnancy: 32.4 yr
Geometric mean
Male partner with
pregnancy result:
1.03 |jg/dL
Male partner without
pregnant result:
1.18 pg/dL
Male reproductive effects:
Fecundity
Women were instructed in
the use of the Clearblue
Easy fertility monitors
consistent with the
manufacturer's guidance
commencing on day six for
tracking daily levels of
E3G and LH. Women also
used the digital Clearblue
Easy home pregnancy test
upon enrollment to ensure
the absence of pregnancy
at study start and on the
day menses was expected
for each cycle under
observation in the study.
Age at outcome:
Average age for males
with pregnancy: 31.6 yr
Average age for males
without pregnancy: 32.4 yr
Cox models for discrete
survival time, which is a
proportional odds
model, adjusted for
age, BMI, cotinine,
parity, serum lipids, and
site (Texas/Michigan)
OR (95% CI), as
fecundability OR
Male only exposure: 0.85
(0.73, 0.99)
Couple exposure:
Female exposure: 1.06
(0.91, 1.24)
Male exposure: 0.82 (0.68,
0.97)
Zhou et al. (2021a) n: 195
China
2018-2019
Cohort
Couples undergoing IVF.
Women with
endometriosis,
hydrosalpinx, abnormal
uterine cavity and men
with azoospermia, severe
oligozoospermia,
asthenospermia and
Blood, other: follicular
fluid, and other: semen
Maternal blood (serum),
follicular fluid, and
seminal plasma from
male partner
Age at Measurement:
Male reproductive effects:
IVF outcome
The IVF outcomes
included were normal
fertilization, good embryo,
blastocyst formation, high-
quality blastocyst,
pregnancy, and live birth
Age at outcome:
Poisson regression
models were adjusted
for age and BMI
RR (95% Cl)c
Normal fertilization
Maternal serum: 0.94 (0.42,
1.93)
Follicular fluid: 0.82 (0.18,
2.39)
Seminal plasma: 1.55 (0.64,
3.3)
Good embryo
8-236
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
dysspermia were excluded Female partner mean
from the study. age: 30.27 yr
Male partner mean age:
31.57 yr
Mean0
Maternal serum:
0.301 [jg/dL
Follicular fluid:
0.742 [jg/dL
Seminal plasma:
0.882 [jg/dL
Median0
Maternal serum:
0.245 [jg/dL
Follicular fluid:
0.178 [jg/dL
Seminal plasma:
0.486 [jg/dL
75th°
Maternal serum:
0.317 [jg/dL
Follicular fluid:
0.326 [jg/dL
Seminal plasma:
1.245 [jg/dL
Female partner mean age:
30.27 yr
Male partner mean age:
31.57 yr
Maternal serum: 1.00 (0.36,
2.38)
Follicular fluid: 0.78 (0.09,
3.03)
Seminal plasma: 1.86 (1.05,
3.11)
Blastocyst formation
Maternal serum: 1.06 (0.2,
3.91)
Follicular fluid: 0.41 (0, 3.63)
Seminal plasma: 1.77 (0.78,
3.58)
High-quality blastocyst
Maternal serum: 1.68 (0.15,
9.43)
Follicular fluid: 0.35 (0, 7.11)
Seminal plasma: 2.66 (0.67,
8)
Pregnancy
Maternal serum: 0.18 (0.01,
1.91)
Follicular fluid: 0.01 (0, 0.03)
Seminal plasma: 0.04 (0,
1.45)
Live birth
Maternal serum: 0.25 (0.01,
2.8)
Follicular fluid: 0 (0, 0.09)
Seminal plasma: 0.01 (0,
1.08)
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study Desfgn Study P°Pulation ExP°sure Assessment Outcome Confounders Effect Esti™£s and 95%
Effects on Morphology or Histology of Male Sex Organs
Huang et al. (2020) Guangxi Birth Cohort
Guangxi
China
July 2015 to
September 2018
Cohort
Study
n: 564 mother-child pairs
Women with singleton
pregnancies that were
included from 8 Maternity
and Child Healthcare
Hospitals in 6 cities of
Guangxi, China
Blood
Maternal blood (serum)
was measured by ICP-
MS
Age at Measurement:
Maternal age at time of
measurement (mean age:
28.76 (SD: 4.66) yr)
Medianb: 0.077 |jg/dL
75thb: 0.123 pg/dL
Quartilesb (pg/dL)
Q1
Q2
Q3
Q4
<0.054
0.055-0.077
0.078-0.123
>0.123
Male reproductive effects:
TV and AGD in infant boys
TV, and AGD-TV
measurements were
undertaken by trained
sonographers using
ultrasonography.
Transverse and
longitudinal grey-scale
images were used to
calculate TV as
tt/6 x length * width * heig
ht. The volumes of both
testes were measured and
an average taken. Two
different measurements of
AGD were obtained using
vernier calipers: the longer
AGD was measured from
the center of the anus to
the cephalad insertion of
the penis (AGDap), and
the shorter AGD was
measured from the center
of the anus to the posterior
base of the scrotum
(AGDas).
Age at outcome:
birth
Multiple linear
(3 (95% Cl)c
regression models were yy
adjusted for BW, GA,
blood sampling time
(mother), alcohol use
pre-pregnancy, BMI,
and age at examination
Q1
Q2
Q3
Q4
Reference
-0.017 (-0.077, 0.043)
-0.024 (-0.085, 0.036)
-0.064 (-0.124, -0.004)
AGDap
Q1
Q2
Q3
Q4
Reference
-0.039 (-0.085, 0.008)
-0.037 (-0.085, 0.010)
-0.060 (-0.110, -0.011)
AGDas
Q1
Q2
Q3
Q4
Reference
-0.020 (-0.091, 0.052)
-0.033 (-0.105, 0.039)
-0.115 (-0.190, -0.039)
AAS = atomic absorption spectrometry; AGD = anogenital distance; AGDap = anopenile distance; AGDas = anoscrotal distance; BMI = body mass index; BW = birth weight;
CI = confidence interval; d = day(s); E2 = estradiol; E3G = estrone-3-glucuronide; fE2 = free estradiol; FSH = follicle stimulating hormone; fT = free testosterone; FSH = follicle
stimulating hormone; GA = gestational age; GFAAS = graphite furnace atomic absorption spectrometry; HTZ = height Z-score; ICP-AES = inductively coupled plasma atomic
emission spectrometry; ICP-MS = inductively coupled plasma mass spectrometry; IVF = in vitro fertilization; LH = luteinizing hormone; LIFE = Longitudinal Investigation of Fertility
and the Environment; LOD = limit of detection; mo = month(s); NHANES = National Health and Nutrition Examination Survey; OR = odds ratio; PIR = poverty-income ratio;
Q = quartile; SA = semen analysis; SBP = systolic blood pressure; SD = standard deviation; SE = standard error; SHBG = sex hormone binding globulin; SPECT = Survey on the
Prevalence in East China for Metabolic Diseases and Risk Factors; T = testosterone; tT = total testosterone; TV = testicular volume; WHO = World Health Organization; yr = year(s).
8-238
-------
Reference and
Study Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bPb measurements were converted from |jg/L to |jg/dL.
°Effects estimates unable to be standardized.
dNo CIs provided.
8-239
-------
Table 8-17
Animal toxicological studies of exposure to Pb and male reproductive effects
Study
Species (Stock/Strain), n, Sex
Timing of
Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
El Shafai et al. (2011) Rat (Wistar)
Control (untreated), M, n = 8
Control (vehicle), M, n = 8
25 mg/kg Pb, M, n = 8
Adulthood (specific
PND NR)
Adult male rats were
dosed via oral gavage for
3 mo. One control group
was not gavaged
(untreated control) and
another control group
was gavaged with vehicle
(vehicle control).
4.26 |jg/dL for
control (untreated)
4.27 |jg/dL for
control (vehicle)
5.27 |jg/dL for
25 mg/kg Pb
Sex Organ
Histopathology
Wang et al. (2013b)
Rat (Sprague-Dawley)
Control (untreated), M, n = 15
0.8/0.3 g/L Pb, M, n = 15
1.5/0.9 g/L Pb, M, n = 15
GD -10 to Dams were dosed via
PND 183 drinking water (0, 0.8, or
1.5 g/L Pb) starting 10 d
prior to mating through
weaning. At weaning 15
males from each group
were dosed via drinking
water to lower levels of
Pb than their dams (0,
0.3, or 0.9 g/L) until 6 mo
of age (approx.
PND 183).
2.65 |jg/dL for
control
18.6 |jg/dL for
0.8/0.3 g/L Pb
55.0 |jg/dL for
1.5/0.9 g/L Pb
Testicular
Weight
8-240
-------
Study
Species (Stock/Strain), n, Sex
Timing of
Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
Wanaetal. (2013a)
Mouse (CD-1)
Control (untreated), M, n = 12
200 ppm Pb, M, n = 12
2000 ppm Pb, M, n = 12
PNDOto PND21
Dams were dosed via
drinking water from
PNDOto 21.
Pups:
PND 22
17.4 |jg/dL for
control
21.2 |jg/dL for
200 ppm Pb
19.1 |jg/L for
2000 ppm Pb
Testosterone
Levels, Sex
Organ
Histopathology,
Accessory Male
Reproductive
Organ Weight,
Testicular
Weight, Semen
Parameters
PND 70
4.40 |jg/dL for
control
3.24 |jg/dL for
200 ppm Pb
5.09 |jg/dL for
2000 ppm Pb
Godinez-Solis et al.
(2019)
Mouse (ICR-CD-1)
Control (untreated), M, n = 4
0.01% Pb, M, n = 6
PND 91 to 136
12 wk old mice were
acclimated for a week
before being dosed via
drinking water for 45 d.
BLL NR for controls
9.4 |jg/dL for
0.01% Pb
Semen
Parameters,
Sperm
Morphology, IVF
8-241
-------
Study
Species (Stock/Strain), n, Sex
Timing of
Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported
(Hg/dL)
Endpoints
Examined
Xie et al. (2020)
Mouse (SPF ICR)
Control (untreated), M, n = 15
50 mg/L Pb, M, n = 15
200 mg/L Pb, M, n = 15
PND 28 to
PND 118
21 d old mice were
acclimated for a week
before being dosed for
90 d via drinking water.
0.602 |jg/dL for
control
6.02 |jg/dL for
50 mg/L Pb
11.8 |jg/dL for
200 mg/L Pb
Semen
Parameters,
Sperm
Morphology,
Sex Organ
Histopathology,
Testicular
Weight,
Accessory Male
Reproductive
Organ Weight
Pavlova et al. (2021)
Mouse (ICR)
Control (vehicle), M, n = 10
80 mg/kg Pb, M, n = 10
PND 60 to 74
60 d old mice were
dosed via oral gavage for
2 wk. Two weeks
following cessation of
exposure, animals were
sacrificed.
1.45 |jg/dL for
control
21.66 |jg/dL for
80 mg/kg Pb
Testicular
Weight, Semen
Parameters, Sex
Organ
Histopathology
BLL = blood lead level; d = day(s); F = female; GD = gestational day; IVF = in vitro fertilization; M = male; mo = month(s); NR = not reported; Pb = lead; PND = postnatal day;
T = testosterone; wk = week(s).
8-242
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United States
Environmental Protection
Agency
EPA/600/R-23/375
January 2024
www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 9: Effects on Other
Organ Systems and Mortality
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
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CONTENTS
DOCUMENT GUIDE 9-iii
LIST OF TABLES 9-vi
LIST OF FIGURES 9-vii
ACRONYMS AND ABBREVIATIONS 9-viii
APPENDIX 9 EFFECTS ON OTHER ORGAN SYSTEMS AND MORTALITY 9-1
9.1 Effects on the Hepatic System 9-2
9.1.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current Review 9-2
9.1.2 Scope 9-2
9.1.3 Epidemiologic Studies on the Hepatic System 9-4
9.1.4 Toxicological Studies on the Hepatic System 9-7
9.1.5 Biological Plausibility 9-8
9.1.6 Summary and Causality Determination 9-11
9.2 Metabolic Effects 9-16
9.2.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current Review 9-16
9.2.2 Scope 9-16
9.2.3 Epidemiologic Studies on Metabolic Effects 9-18
9.2.4 Toxicological Studies on Metabolic Effects 9-23
9.2.5 Summary and Causality Determination 9-24
9.3 Effects on the Gastrointestinal System 9-25
9.3.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current Review 9-25
9.3.2 Scope 9-25
9.3.3 Epidemiologic Studies on the Gastrointestinal System 9-27
9.3.4 Toxicological Studies on the Gastrointestinal System 9-27
9.3.5 Summary and Causality Determination 9-27
9.4 Effects on the Endocrine System 9-28
9.4.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current Review 9-28
9.4.2 Scope 9-28
9.4.3 Epidemiologic Studies on the Endocrine System 9-30
9.4.4 Toxicological Studies on the Endocrine System 9-32
9.4.5 Summary and Causality Determination 9-32
9.5 Effects on the Musculoskeletal System 9-34
9.5.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current Review 9-34
9.5.2 Scope 9-35
9.5.3 Epidemiologic Studies on the Musculoskeletal System 9-36
9.5.4 Toxicological Studies on the Musculoskeletal System 9-40
9.5.5 Biological Plausibility 9-41
9.5.6 Summary and Causality Determination 9-45
9.6 Effects on Ocular Health 9-50
9.6.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current Review 9-50
9.6.2 Scope 9-50
9.6.3 Epidemiologic Studies on Ocular Health 9-52
9.6.4 Toxicological Studies on Ocular Health 9-53
9.6.5 Summary and Causality Determination 9-54
9.7 Effects on the Respiratory System 9-55
9-iv
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9.7.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current Review 9-55
9.7.2 Scope 9-55
9.7.3 Epidemiologic Studies on the Respiratory System 9-57
9.7.4 Toxicological Studies on the Respiratory System 9-59
9.7.5 Summary and Causality Determination 9-60
9.8 Mortality 9-61
9.8.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current Review 9-61
9.8.2 Scope 9-62
9.8.3 Total (non-Accidental) Mortality 9-63
9.8.4 Cause-Specific Mortality 9-68
9.8.5 Biological Plausibility 9-69
9.8.6 Summary and Causality Determination 9-70
9.9 Evidence Inventories - Data Tables to Summarize Study Details 9-74
9.10 References 9-149
9-v
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LIST OF TABLES
Table 9-1 Evidence that is suggestive of, but not sufficient to infer, a causal relationship between Pb
exposure and hepatic effects 9-14
Table 9-2 Summary of evidence for a likely to be causal relationship between Pb exposure and
musculoskeletal effects 9-48
Table 9-3 Summary of evidence for a causal relationship between Pb exposure and total mortality 9-73
Table 9-4 Epidemiologic studies of exposure to Pb and hepatic effects 9-74
Table 9-5 Animal toxicological studies of exposure to Pb and hepatic effects 9-82
Table 9-6 Epidemiologic studies of exposure to Pb and metabolic effects 9-86
Table 9-7 Animal toxicological studies of exposure to Pb and metabolic effects 9-100
Table 9-8 Animal toxicological studies of exposure to Pb and gastrointestinal effects 9-102
Table 9-9 Epidemiologic studies of exposure to Pb and endocrine effects 9-103
Table 9-10 Animal toxicological studies of exposure to Pb and endocrine effects 9-112
Table 9-11 Epidemiologic studies of exposure to Pb and musculoskeletal effects 9-115
Table 9-12 Animal toxicological studies of exposure to Pb and musculoskeletal effects 9-127
Table 9-13 Epidemiologic studies of exposure to Pb and ocular effects 9-128
Table 9-14 Animal toxicological studies of Pb exposure and ocular effects 9-133
Table 9-15 Epidemiologic studies of Pb exposure and respiratory effects 9-134
Table 9-16 Animal toxicological studies of exposure to Pb and respiratory effects 9-139
Table 9-17 Epidemiologic studies of Pb exposure and total mortality 9-141
9-vi
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LIST OF FIGURES
Figure 9-1 Potential biological pathways for hepatic effects following exposure to Pb. 9-11
Figure 9-2 Potential biological pathways for musculoskeletal effects following exposure to Pb. 9-44
Figure 9-3 Effect estimates for associations of blood Pb with all-cause mortality. 9-64
Figure 9-4 Dose-response relationship between blood Pb levels and all-cause mortality. 9-65
9-vii
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ACRONYMS AND ABBREVIATIONS
AAS atomic absorption spectrometry
AD Alzheimer's disease
ALAD S-aminolevulenic acid dehydratase
ALP alkaline phosphatase
ALT alanine aminotransferase
AMD age-related macular degeneration
AOPP advanced oxidation protein products
AQCD Air Quality Criteria Document
ARCA Automobile Racing Clube of America
AST aspartate aminotransferase
AV/TV adipocyte volume/total volume
BLL blood lead level
BMD bone mineral density
BMI body mass index
BMP bone morphogenic protein
BV/TV bone volume to total volume
C2C serum cleavage neoepitope of type II
collagen
Ca2+ calcium ion(s)
CAT catalase
C-R concentration-response
CAR Cortisol awakening response
Cd cadmium
CD control diet
CHEER Children's Health and Environmental
Research
CHF congestive heart failure
CI confidence interval
CK18 cytokeratin 18
COMP cartilage oligomeric matrix protein
CPU carboxypropeptide of type II collagen
CRP C-reactive protein
CVD cardiovascular disease
CYP Cytochrome P450
d day(s)
DBP diastolic blood pressure
DMFS Delayed, missing, and filled surfaces
DMFT decayed, missing, and filled teeth
DXA Dual-energy X-ray absorptiometry
ECRHS European Community Respiratory
Health Survey
EDTA ethylenediaminetetraacetic acid
EGF epidermal growth factor
eGFR estimated glomerular filtration rate
ELEMENT Early Life Exposure in Mexico to
Environmental Toxicants
ER endoplasmic reticulum
ERSD end-stage renal disease
F# filial generation
FBG fasting blood glucose
FEV1 forced expiratory volume in one second
FIB-4 fibrosis-4
FT3 free triiodothyronine
FT4 free thyroxine
FVC forced vital capacity
GADA glutamic acid decarboxylase antibodies
GD gestational day
GDM gestational diabetes mellitus
GDS Gesell Developmental Schedules
GFAAS graphite furnace atomic absorption
spectrometry
GFR glomerular filtration rate
GGT gamma-glutamyl transferase
GH growth hormone
GI gastrointestinal
GM geometric mean
GPx glutathione peroxidase
GSH glutathione
GSH-PX glutathione peroxidase
Hb hemoglobin
HDL high-density lipoprotein
HDL-C high-density lipoprotein cholesterol
HF hepatic fibrosis
HOMA- p HOMA of P-cell function
HOMA-IR Homeostatic Model Assessment for
Insulin Resistance
HR hazard ratio
HS hepatic steatosis
ICP-MS inductively coupled plasma mass
spectrometry
ID iron deficient
IHC immunohistochemistry
IHD ischemic heart disease
i.p. intraperitoneal
IOP intraocular pressure
ISA Integrated Science Assessment
KARE Korean Association Resource
KNHANES Korea National Health and Nutrition
Examination Survey
K-XRF K-shell X-ray fluorescence
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LDL low-density lipoprotein
LDL-C low-density lipoprotein cholesterol
LOD limit of detection
mo month(s)
MDA malondialdehyde
MetS metabolic syndrome
METS Modeling the Epidemiologic Transition
Study
MI myocardial infarction
microCT micro-computed tomography
mRNA messenger ribonucleic acid
NAAQS National Ambient Air Quality
Standards
NAFLD nonalcoholic fatty liver disease
NANC noncholinergic
NAS Normative Aging Study
NASCAR National Association for Stock Car
Auto Racing
NHANES National Health and Nutrition
Examination Survey
NF -kB nuclear factor kappa B
NP nanoparticle
OA osteoarthritis
OLD obstructive lung disease
OLF obstructive lung function
OR odds ratio
Pb lead
PbO lead oxide
PCNA proliferating cell nuclear antigen
PCR polymerase chain reaction
PD Potential difference
PECOS Population, Exposure, Comparison,
Outcome, and Study Design
PSS perceived stress score
PIR poverty-income ratio
PM particulate matter
PND postnatal day
PROGRESS Programming Research in Obesity,
Growth, Environment and Social
Stressors
PTE! parathyroid hormone
PTElrP parathyroid hormone-related protein
qRT-PCR real-time quantitative reverse
transcription-polymerase chain reaction
RBC red blood cell
RCT randomized control trial
RR relative risk
RT-PCR reverse transcription-polymerase chain
reaction
SBP systolic blood pressure
SBEE1C Shiwha and Banwol Environmental
Elealth Cohort
SD standard deviation
SE standard error
SES socioeconomic status
SNP single nucleotide polymorphism
SOD superoxide dismutase
SPECT single photon emission computed
tomography
SSBI sugar sweetened beverage intake
T-SOD total superoxide dismutase
T# tertile #
TACT Trial to Assess Chelation Therapy
TB total bilirubin
TBARS thiobarbituric acid reactive substance
TC total cholesterol
TEM transmission electron microscopy
Tg thyroglobulin
TGAb thyroglobulin antibody
TGF-pi transforming growth factor-beta 1
TNF tumor necrosis factor
TRI Toxics Release Inventory
TSH thyroid-stimulating hormone
TPOAb thyroid peroxidase antibody
Q quartile
wk week(s)
yr year(s)
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APPENDIX 9 EFFECTS ON OTHER ORGAN SYSTEMS
AND MORTALITY
Summary of Causality Determinations for Pb Exposure and Effects on Other Organ Systems and
Mortality
This appendix characterizes the scientific evidence that supports causality determinations for
lead (Pb) exposure and hepatic effects, metabolic effects, gastrointestinal effects, endocrine system
effects, effects on bone and teeth, effects on ocular health, and respiratory effects. The types of studies
evaluated within this appendix are consistent with the overall scope of the ISA as detailed in the
Process Appendix (see Section 12.4). In assessing the overall evidence, strengths and limitations of
individual studies were evaluated based on scientific considerations detailed in the Table 12-5 of the
Process Appendix (Section 12.6.1). More details on the causal framework used to reach these
conclusions are included in the Preamble to the ISA (U.S. EPA. 2015). The evidence presented
throughout this appendix supports the following causality conclusions:
Outcome Group
Causality Determination
Hepatic Effects
Suggestive of, but not sufficient to infer, a causal
relationship
Metabolic Effects
Inadequate
Gastrointestinal Effects
Inadequate
Endocrine System Effects
Inadequate
Musculoskeletal Effects
Likely to be Causal
Ocular Health Effects
Inadequate
Respiratory Effects in
Populations without Asthma
Inadequate
Total (Nonaccidental) Mortality
Causal
The Executive Summary, Integrated Synthesis, and all other appendices of the 2024 Pb ISA can be
found at https://assessments.epa.gov/isa/document/&deid=359536.
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9.1
Effects on the Hepatic System
9.1.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current
Review
The 2013 Integrated Science Assessment for Lead (hereinafter referred to as the 2013 Pb ISA)
concluded that "because of the insufficient quality of studies, the available evidence was inadequate to
determine if there is a causal relationship between Pb exposure and hepatic effects" (U.S. EPA. 2013).
Epidemiologic evidence from a limited number of occupational studies demonstrated impaired liver
function in Pb-exposed workers. However, the internal validity and generalizability of these studies was
limited by cross-sectional study designs, lack of consideration for potential confounders, and notably
higher blood Pb levels (BLLs) (>29 (ig/dL) than the general population. Similarly, toxicological studies
observed changes in liver function enzymes and other markers of liver health in animals exposed to Pb,
but the use of bolus injections as a common route of exposure and high BLLs (>30 (ig/dL) introduced
uncertainty regarding their relevance to human exposures.
9.1.2 Scope
The scope of this section is defined by Population, Exposure, Comparison, Outcome, and Study
Design (PECOS) statements. The PECOS statement defines the objectives of the review and establishes
study inclusion criteria, thereby facilitating identification of the most relevant literature to inform the Pb
ISA.1 In order to identify the most relevant literature, the body of evidence from the 2013 Pb ISA was
considered in the development of the PECOS statements for this Appendix. Specifically, well-established
areas of research; gaps in the literature; and inherent uncertainties in specific populations, exposure
metrics, comparison groups, and study designs identified in the 2013 Pb ISA inform the scope of this
Appendix. The 2013 Pb ISA used different inclusion criteria than the 2024 Pb ISA, and the studies
referenced therein often do not meet the current PECOS criteria (e.g., due to higher or unreported
biomarker levels). Studies included in the 2013 Pb ISA, including many that do not meet the current
PECOS criteria, are discussed in this appendix to establish the state of the evidence prior to this
assessment. Except for supporting evidence used to demonstrate the biological plausibility of Pb-
associated effects on the hepatic system, recent studies were only included if they satisfied all the
components of the following discipline-specific PECOS statements:
'The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
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Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect.
Exposure: Exposure to Pb2 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure;3 or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles).
Outcome: Effects on the hepatic system.
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages).
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a BLL of 30 (ig/dL or below.4,5
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control.
Outcomes: Effects on the hepatic system.
Study design: Controlled exposure studies of animals in vivo.
2Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area of particular
relevance to the National Ambient Air Quality Standards (NAAQS) review (e.g., longitudinal studies designed to
examine recent versus historical Pb exposure).
3Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 |im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 |im3 (PM2.5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is difficult
to assess the representativeness of these concentrations to population exposure (Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PNLs with blood Pb levels (BLLs) are lacking.
4Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
5This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLL.
The 95th percentile of the 2011-2016 National Health and Nutrition Examination Survey (NHANES) distribution of
BLL in children (1-5 years; n = 2,321) is 2.66 (ig/dL (Eganet al.. 2021) and the proportion of individuals with BLL
that exceed this concentration varies depending on factors including (but not limited to) housing age, geographic
region, and a child's age, sex, and nutritional status.
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9.1.3 Epidemiologic Studies on the Hepatic System
Epidemiologic evidence evaluated in the 2013 Pb ISA (U.S. EPA. 2013) was limited to a small
number of occupational studies that demonstrated impaired liver function in Pb-exposed workers.
However, the internal validity and generalizability of these studies was limited by cross-sectional study
designs, lack of consideration for potential confounders, and notably higher BLLs (>29 (ig/dL) than the
general population. Recent epidemiologic studies of the hepatic system generally examine one of three
groups of endpoints: (1) direct evaluation of liver injury (e.g., nonalcoholic fatty liver disease [NAFLD]
and hepatic fibrosis); (2) serum biomarkers of liver function (e.g., alanine aminotransferase [ALT],
aspartate aminotransferase [AST], alkaline phosphatase [ALP], and gamma-glutamyl transferase [GGT]);
and (3) serum lipids (e.g., fatty acids, lipids, and cholesterol). Results from recent studies provide
inconsistent evidence of an association between BLLs and direct or indirect measures of liver damage.
Recent studies evaluating hepatic effects are generally limited to cross-sectional analyses, which are
unable to establish temporality between exposure and outcome. Additionally, with biomarkers of Pb
exposure, it is difficult to characterize the specific timing, duration, frequency, and level of Pb exposure
that contributed to associations observed with liver function. This uncertainty may apply particularly to
assessments of BLLs, which in nonoccupationally-exposed adults, reflect both current exposures and
cumulative Pb stores in bone that are mobilized during bone remodeling. Measures of central tendency for
Pb biomarker levels used in each study, along with other study-specific details, including study
population characteristics and select effect estimates, are highlighted in Table 9-4. An overview of the
recent evidence is provided below.
9.1.3.1 Direct Evaluation of Liver Injury
A limited number of recent cross-sectional studies examined the association between BLLs and
liver injury, including NAFLD and fibrosis (Chung et al.. 2020; Reiaet al.. 2020; Werder et al.. 2020;
Zhai et al.. 2017). These studies, which use a variety of diagnostic tools, provide inconsistent evidence of
an association between BLLs and NAFLD and fibrosis. Liver biopsy is the gold standard for evaluating
NAFLD and liver fibrosis, but it is an invasive and cost prohibitive procedure. Therefore, epidemiologic
studies often rely on alternative measurement techniques, including imaging, biomarkers, and biomarker-
based prediction models. Imaging—either ultrasonic or magnetic resonance—generally has greater
sensitivity and specificity than reliance on biomarkers.
A recent cross-sectional study of adults in the Yangtze River Delta in China examined the
relationship between BLLs and NAFLD measured by ultrasound (Zhai et al.. 2017). In addition to using
ultrasonic imaging, this study included a large number of participants (n = 2,011). In sex-stratified
models, Zhai et al. (2017) reported higher odds of NAFLD associated with higher BLL quartiles after
adjusting for a range of demographic and hepatic and metabolic health factors. The observed associations
were stronger in magnitude among men (odds ratio [OR] = 2.168 [95% CI: 0.989, 4.751] quartile 4 versus
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quartile 1) compared with women (OR= 1.613 [95% CI: 1.082, 2.405] quartile 4 versus quartile 1);
however, the effect estimates in men were much less precise due to a smaller sample of men in the study
population. Given the imprecise estimates for men (i.e., wide 95% Cis), it is difficult to draw conclusions
on sex-specific comparisons.
Results from other recent cross-sectional studies are inconsistent. In a small exploratory analysis
of oil spill response workers with low BLLs (mean = 1.82 (.ig/dL). Werder et al. (2020) evaluated the
association between BLLs and cytokeratin 18 (CK18), a serologic biomarker of hepatocyte death that has
been used as a marker forNAFLD. The authors observed an association between BLLs and caspase-
cleaved fragment CK18 (CK18 M30), but not whole protein CK18 (CK18 M65). Notably, CK18 M65 has
performed better as a measure of NAFLD than CK18 M30 (Lee et al.. 2020). adding further ambiguity to
the observed results. Additionally, Werder et al. (2020) examined a range of heavy metals and markers of
inflammation and did not adjust for multiple testing, which increases the likelihood of chance findings
and may explain the inconsistent results. In addition to this weak evidence of an association between
BLLs and markers of NAFLD, (Chung et al.. 2020) analyzed data from the Korea National Health and
Nutrition Examination Survey (KNHANES) and reported null or negative sex-specific associations
between BLLs and scores on the Hepatic Steatosis Index, a validated biomarker-based prediction model
of NAFLD. The authors also observed negative associations between BLLs and Fibrosis 4 Index, a
similarly validated model for fibrosis. This larger analysis (n = 4,420) reported similar mean BLLs
(1.81 (ig/dL) as those reported in Werder et al. (2020).
In addition to studies examining NAFLD and fibrosis separately, (Reia et al.. 2020) used a
biomarker-based index to estimate fibrosis level in National Health and Nutrition Examination Survey
(NHANES) participants with NAFLD. In this case, fibrosis level was used as an indicator of NAFLD
severity. Reia et al. (2020) reported large, but imprecise associations between BLL quartiles and
advanced liver fibrosis. For example, the authors noted that participants in the highest quartile of BLLs
(>1.62 (ig/dL) had nearly 5-fold higher odds of advanced liver fibrosis (OR = 4.93 [95% CI: 1.88, 11.24])
compared to participants in the lowest quartile (<0.64 (ig/dL). Despite having a large sample size, the
authors only examined severe liver fibrosis, which likely resulted in a small number of cases (total cases
not reported) and would have decreased the statistical power of the study. Limited statistical power
resulting from a small sample size reduces the likelihood of detecting a true effect.
9.1.3.2 Serum Biomarkers of Liver Function
Serum biomarkers can be used as indirect evidence of liver damage. For example, elevated levels
of ALT or AST can indicate the presence of necrosis in the liver, and elevated levels of bilirubin, ALP, or
GGT can be associated with cholestasis. However, changes in serum biomarker levels are also related to
effects on other biological systems. Elevated GGT can also occur with chronic heart failure, and elevated
ALP can be used to detect bone disorders. Therefore, studies evaluating these biomarkers in combination
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are more likely to provide evidence of abnormal liver function relative to studies evaluating a single
biomarker.
There have been a limited number of recent epidemiologic studies that evaluated serum
biomarkers of liver function, including a longitudinal study (Pollack et al.. 2015) and a few cross-
sectional analyses (Chen et al.. 2019; Obcng-Gvasi. 2019; Christensen et al.. 2013). Recent studies, which
adjust for a wide range of potential confounders, provide some evidence of an association between BLLs
and serum biomarkers, but results are not entirely consistent, and the implications of some associations
are unclear. Specifically, a small prospective cohort study of premenopausal women evaluated the percent
change in AST, ALT, ALP, and bilirubin over the course of an 8-week follow-up (Pollack et al.. 2015).
The authors reported imprecise positive associations between AST (5.02% [95% CI: -1.36%, 11.41%]),
ALT (6.39% [95% CI: 3.07%, 9.72%]), and ALP (2.14% [95% CI: -5.05%, 9.33%]) and BLLs measured
at baseline (mean = 1.03 (ig/dL), but no association between bilirubin and BLLs (-0.20% [95% CI:
-7.50%, 7.11%]). The clinical relevance of these findings is uncertain given the majority of the study
population fell well within the normal ranges of each of the biomarkers. While small shifts in biomarker
levels may have important public health implications, the importance of these findings would be better
substantiated with evidence of associations between Pb exposure and more direct measures of liver injury.
A recent cross-sectional study of adults living near an e-waste facility in China better addresses clinical
relevance by examining the association between BLLs and abnormal liver function, defined as having two
or more transaminases (AST, ALT, GGT) elevated above the normal range, or having one transaminase at
least twice as high as the upper bound of the normal range (Chen et al.. 2019). In this study, which had
notably higher median BLLs (5.1 to 8.7 (ig/dL across study locations), BLLs were associated with a large,
but imprecise increase in the odds of abnormal liver function (OR = 1.94 [95% CI: 1.00, 3.73] per
1 (ig/dL higher BLL).
Results from recent large cross-sectional NHANES analyses examining a single serum biomarker
of liver function were inconsistent. In an analysis of 2003-2004 NHANES participant's ages 12 years and
older, Christensen et al. (2013) reported null associations between BLL quartiles and ALT levels. An
analysis restricted to adult participants of more recent NHANES survey cycles (2011-2016) observed
higher odds of GGT levels above the study population median (18 U/L) associated with each 1 (ig/dL
higher BLL (OR= 1.94 [95% CI: 1.652, 2.28] for young adults and 1.34 [95% CI: 1.14, 1.58] for middle-
aged adults) (Obeng-Gvasi. 2019). Similar to the Pollack et al. (2015) study, the median GGT levels in
this study were within the normal range.
9.1.3.3 Serum Lipids
Many fatty acids, lipids, and cholesterol are synthesized and eliminated in the liver; the
relationships among them and their relevance to other aspects of human health, including metabolic
effects (Section 9.2) and cardiovascular effects (Appendix 4). are complex. Although increases or
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decreases in serum or liver cholesterol levels may be associated with liver damage, it can be challenging
to determine whether the changes are a consequence of said damage or a contributing factor in disease
progression (Argucllo et al.. 2015; Chrostek et al.. 2014). Recent epidemiologic studies of serum lipids
have been conducted in populations of adults and children and include a mix of prospective cohorts and
cross-sectional designs. Recent studies also account for a range of potential confounders, including
demographics and socioeconomic status (SES) factors, medical history, and medication use. Associations
between BLLs and serum lipids have been largely inconsistent across both lifestages.
In a recent study including a subset of the Veterans Affairs Normative Aging Study (NAS) cohort
with healthy older adults, Peters et al. (2012) examined the associations between BLLs at baseline and
serum lipid levels after three to four years of follow-up. The authors reported higher odds of clinically
elevated total cholesterol associated with higher BLLs (OR= 1.08 [95%: 0.99, 1.19] per 1 (.ig/dL higher
BLL). Associations with clinical cut points for other serum lipids were either null (elevated triglycerides
and low-density lipoprotein cholesterol [LDL] cholesterol) or negative (low high-density lipoprotein
[HDL] cholesterol). Cross-sectional studies of adult populations, including analyses of nationally
representative health survey data (Xu et al.. 2021; Lee and Kim. 2016) and a small analysis of adults of
African descent (Ettinger et al.. 2014). are also inconsistent. Results across these studies (see Table 9-2)
provide no discernable pattern of associations between BLLs and triglycerides, LDL cholesterol, or HDL
cholesterol. BLL measures of central tendency were low across the evaluated studies (<5 (ig/dL) and do
not appear to explain the inconsistencies.
Results from studies in children are similarly inconsistent. Two recent studies of serum lipids
analyzed data from separate birth cohorts in Mexico - the Early Life Exposure in Mexico to
Environmental Toxicants (ELEMENT) study (Liu et al.. 2020) and the Programming Research in
Obesity, Growth, Environment and Social Stressors (PROGRESS) birth study (Kupsco et al.. 2019). In
children ages 4 to 6, Kupsco et al. (2019) reported null associations between prenatal BLLs and serum
triglycerides and non-HDL cholesterol. In contrast, in an analysis including older children and teens, Liu
et al. (2020) observed higher triglyceride Z-scores in children with prenatal BLLs >5 (ig/dL compared to
those with BLLs less than 5 (ig/dL (0.58 [95% CI: -0.05, 1.20]). The authors observed negative
associations between prenatal BLLs and cholesterol Z-scores (total, LDL, and HDL). A large cross-
sectional analysis of NHANES participants ages 12 to 19 noted 2.3% (95% CI: 0.3%, 4.2%) higher LDL
cholesterol levels and 0.6% (95% CI: -0.1%, 1.3%) higher total cholesterol levels per 1 (ig/dL higher
BLL (Xu et al.. 2017). The authors observed null (total cholesterol and HDL cholesterol) or negative
(triglycerides) associations between BLLs and other serum lipids.
9.1.4 Toxicological Studies on the Hepatic System
As described in the 2013 Pb ISA, evidence from toxicological studies indicates exposure to Pb
can result in altered liver function and hepatic oxidative stress (U.S. EPA. 2013). A few studies reported
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Pb-induced decreases in cytochrome P450 (CYP) enzymes (Phase I xenobiotic metabolism), as well as
Pb-induced decreases in serum protein and albumin levels and increased AST, ALT, ALP, and GGT
activities (indicators of decreased liver function), increased oxidative stress, and decreased antioxidant
status. A number of recent studies have corroborated findings of Pb exposure and decreased liver function
(Barkaoui et al.. 2020; Dumkova et al.. 2020b; Gao et al.. 2020; Andielkovic et al.. 2019; Laamech et al..
2017; Long et al.. 2016; Liu et al.. 2013; Berrahal et al.. 2011). While impaired lipid metabolism was
reported in the 2013 Pb ISA, results from recent studies of cholesterol have been inconsistent. Laamech et
al. (2017) found an increase in total cholesterol in mice given Pb acetate in their drinking water (BLL:
18 (ig/dL). Conversely, Dumkova et al. (2020a) found lower levels of total cholesterol in rats that were
given Pb oxide nanoparticles by inhalation (BLLs: 3.1-8.5 (.ig/dL); however, the latter group did report an
increase in lipid droplets by liver histology [BLLs: 3.1-17.8 (ig/dL; (Dumkova et al.. 2020a; Dumkova et
al.. 2020b; Dumkova et al.. 2017)1. Observation of Pb-associated increases in hepatic oxidative stress, as
indicated by a decrease in glutathione (GSH) levels and catalase (CAT), superoxide dismutase (SOD),
and glutathione peroxidase (GPx) activities has been found in additional recent studies of oral Pb
exposure [drinking water: 21.4-29.0 (ig/dL (Barkaoui et al.. 2020; Andielkovic et al.. 2019; Long et al..
2016); oral gavage: 18.5-30.2 (ig/dL (Gao et al.. 2020; Laamech et al.. 2017; Li et al.. 2017)1.
Since the 2013 Pb ISA, several recent studies have reported perturbations related to oxidative
stress in addition to the endpoints noted above. For example, Andielkovic et al. (2019) found changes in
multiple parameters of oxidative stress in liver and kidney tissue in male rats, indicative of an oxidative
stress response to Pb exposure (BLL: 29.0 (.ig/dL). Long et al. (2016) also reported several markers of
oxidative damage and response, in mouse liver tissue. They showed in addition, consistent with an
oxidative damage response, attenuation of such response after administration of proanthocyanidins, which
are naturally occurring antioxidant compounds. The same authors reported changes in several markers
that are consistent with a generalized endoplasmic reticulum (ER) response in the liver to environmental
stressors. Likewise, Liu et al. (2013) showed Pb responsiveness of ER stress markers, and the antagonistic
effect of quercetin (a natural flavonoid) on this response. Barkaoui et al. (2020) reported finding
alleviation of Pb-induced oxidative effects from administration of antioxidative, phenolic compounds
extracted from Plantigo albicans.
Cell death by apoptosis may be a downstream result of the molecular sequelae of Pb exposure
described in the preceding paragraph. Indeed, such a result has been reported in mouse livers, both
phenotypically and via molecular markers (Dumkova et al.. 2017; Long et al.. 2016).
9.1.5 Biological Plausibility
This section describes biological pathways that potentially underlie effects of Pb on the liver and
hepatic function. Figure 9-1 depicts the proposed pathways as a continuum of upstream events, connected
by arrows, which may lead to downstream events observed in epidemiologic studies. This discussion of
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how exposure to Pb may lead to hepatic effects contributes to an understanding of the biological
plausibility of epidemiologic results evaluated above. Note that the structure of the biological plausibility
sections and the role of biological plausibility in contributing to the weight-of-evidence analysis used in
the 2024 Pb ISA are discussed in Section IS.4.2.
The hepatic effects of Pb exposure have been studied in many experimental models. The pathway
proposed, outlined in Figure 9-1, involves the induction of oxidative stress and inflammation leading to
downstream cellular loss and metabolic changes that could plausibly be responsible for the development
of health effects in the liver. Oxidative stress control and inflammation are highly regulated processes and
are tightly linked. As discussed above and in both the 2013 Pb ISA and 2006 Pb Air Quality Criteria
Document (AQCD), inflammatory signaling and marker of oxidative stress have been found in the livers
of animals exposed to Pb (see Section 9.1.3 and (U.S. EPA. 2013. 2006). Hepatic inflammation and
oxidative stress co-occur thus it is difficult to determine if one process precedes the other, thus, they are
grouped in the same grey box in Figure 9-1.
Regulation of inflammation and oxidative stress involve widespread gene expression changes that
could plausibly alter the expression of metabolizing enzymes and proteins necessary for cholesterol
synthesis and maintaining lipid homeostasis which could lead to fat accumulation and subsequent fatty
liver disease. As discussed in the 2013 Pb ISA, Pb treatment can cause elevated cholesterol levels through
changes in cholesterol synthesis pathways in the liver. Pb can also alter the expression and activity of
CYP enzymes that are important in the response to xenobiotics as well as metabolism of cholesterol-
derived steroid hormones. A recent study in knockout mice showed that mice deficient in the 11-1
inflammatory mediators were protected from the hypercholesterolemia in response to Pb compared to
wild type mice (Koiima et al.. 2012). Knockout mice also did not experience the messenger ribonucleic
acid (mRNA) upregulation cholesterol synthesizing enzymes HMGR and Cvp51 or the downregulation of
bile acid synthesizing enzyme Cyp7al. These data support the necessity of inflammation to the regulation
of cholesterol metabolism and are the basis for the solid line from inflammation to the box containing
CYP activity and altered cholesterol synthesis in Figure 9-1.
Excessive damage from oxidative stress and inflammatory responses could lead to cell death
which, in excess, could lead to changes in hepatocyte structure and ultimately decrease liver function. As
discussed above and in the 2013 Pb ISA and 2006 Pb AQCD, many animal studies have shown that Pb
exposure of varying durations and developmental stages results in liver injury, which is most commonly
measured as increased activity of liver enzymes (e.g., AST, ALT, ALP) in the blood serum or plasma.
Increases of liver enzyme activity have been seen in the serum of humans occupationally exposed to Pb
(Mazumdar and Goswami. 2014; U.S. EPA. 2013). As mentioned above, elevated liver enzymes in the
blood can serve as an indirect markers of liver damage. Previous research has shown that exposure to Pb
in animal models can lead to upregulation of cell death pathways (U.S. EPA. 2013) and more recent
studies provide additional support (Almasmoum et al.. 2019; Abu-Khudir et al.. 2017; Hasanein et al..
2016; Long et al.. 2016; Mabrouk et al.. 2016; Liu et al.. 2013; Pal et al.. 2013; Liu et al.. 2012. 2011).
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Studies have shown that treatment with antioxidants, like vitamin E (Almasmoum et al.. 2019). vitamin C
(Upadhvav et al.. 2009). or therapeutic compounds that have anti-inflammatory and antioxidant properties
(Abu-Khudir et al.. 2017; Hasanein et al.. 2016; Long et al.. 2016; Mabrouk et al.. 2016; Liu et al.. 2013;
Pal et al.. 2013; Liu et al.. 2012) can prevent the Pb-induced upregulation of apoptotic pathways and
concomitantly reduced both markers of oxidative damage and serum markers of liver injury. Interestingly,
some therapeutic compounds reduce the liver Pb burden suggesting that the reduction in oxidative stress
may be caused by toxicokinetic changes that reduce the liver Pb exposure concentration (Liu et al.. 2013;
Liu et al.. 2011). however, some studies have seen that antioxidant treatment can reduce oxidative stress
even while live Pb levels remain elevated suggesting that oxidative stress is directly related to
downstream liver damage (Almasmoum et al.. 2019; Long et al.. 2016; Mabrouk et al.. 2016; Reckziegel
et al.. 2016). Together these data provide support for the solid line from the box containing inflammation
and oxidative stress to cell death.
Excessive cell loss can result in changes to liver architecture and trigger repair processes that can
lead to liver scarring, both of which can lead to loss of liver function. The 2013 Pb ISA discussed studies
that showed that Pb treatment led to noticeable histologic changes including signs of increased fibrotic
liver changes (U.S. EPA. 2013). More recent work supports this with evidence that liver histologic
changes are accompanied by increased markers of apoptosis and necrosis (Long et al.. 2016; Mabrouk et
al.. 2016). A study also showed that 4 months of Pb exposure in rats increased wound repair signaling
pathways which corresponded to increased deposition of extracellular matrix proteins in the liver (Perez
Aguilar et al.. 2014). Sufficient damage to the liver can reduce liver function which can be measured as a
reduced level of protein in the blood. Recent studies have shown decreases in serum proteins following
Pb exposure that coincide with molecular or histological signs of liver damage (Almasmoum et al.. 2019;
El-Tantawv. 2016; Hasanein et al.. 2016). Similar evidence is seen in the 2013 Pb ISA. Together, it is
plausible that widespread cell death in the liver can lead to changes in hepatocyte structure that leads to
liver damage and resulting decline in liver function.
The proposed pathway leading from Pb exposure to hepatic health effects begins with the
induction of inflammation and increase in oxidative stress. This results in both changes in metabolizing
enzymes and cholesterol synthesis that could be responsible for fatty accumulation in the liver.
Widespread oxidative damage results in cell loss which could disrupt the normal liver structure and
contribute to loss of liver function. Together, the evidence supports a plausible pathway from Pb exposure
to the hepatic effects seen in epidemiologic and animal tox studies.
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1
Decreased liver function
Liver injury
Altered cholesterol synthesis
1
¦
Altered CYP activity
Fatty liver disease
CYP = cytochrome P450.
Note: The boxes above represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and
the arrows indicate a proposed relationship between those effects. Solid arrows denote evidence of essentiality as provided, for
example, by an inhibitor of the pathway used in an experimental study involving Pb exposure. Dotted arrows denote a possible
relationship between effects. Shading around multiple boxes is used to denote a grouping of these effects. Arrows may connect
individual boxes, groupings of boxes, and individual boxes within groupings of boxes. Progression of effects is generally depicted
from left to right and color coded (white, exposure; green, initial effect; blue, intermediate effect; orange, effect at the population
level or a key clinical effect). Here, population-level effects generally reflect results of epidemiologic studies. When there are gaps in
the evidence, there are complementary gaps in the figure and the accompanying text below. The structure of the biological
plausibility sections and the role of biological plausibility in contributing to the weight-of-evidence analysis used in the 2022 Pb ISA
are discussed in IS.7.2.
Figure 9-1 Potential biological pathways for hepatic effects following
exposure to Pb.
9.1.6 Summary and Causality Determination
The 2013 Pb ISA (U.S. EPA. 2013) concluded that the available evidence was "inadequate to
determine if there is a causal relationship between Pb exposure and hepatic effects." A limited number of
occupational epidemiologic studies evaluated potential associations between higher BLLs and lower
serum protein and albumin levels and higher liver function enzymes, oxidative stress, and antioxidant
status. The implications of the occupational epidemiologic evidence were limited because of the cross-
sectional design of the studies, the high BLLs examined (means >29 |ig/dL). and the lack of consideration
for potential confounding by factors such as age, diet, BMI, smoking, or other occupational exposures.
Similar changes in liver function enzymes were found in mature animals exposed to high levels of Pb
during adulthood, and animals exposed during gestation and lactation. Pb exposure was also shown to
impair lipid metabolism in animals, as evidenced by increased hepatic cholesterogenesis, and altered
triglyceride and phospholipid levels (Sharma et al.. 2010; Ademuviwa et al.. 2009; Khotimchenko and
Kolenchenko. 2007). Multiple toxicological studies observed Pb-related increases in hepatic oxidative
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stress, generally indicated by an increase in lipid peroxidation along with a decrease in GSH levels and
CAT, SOD, and GPx activities (Pandva et al.. 2010; Sharma et al.. 2010; Yu et al.. 2008; Adcgbcsan and
Adenuga. 2007; Jurczuk et al.. 2007; Khotimchenko and Kolenchenko. 2007; Jurczuk et al.. 2006).
However, the relevance of the toxicological evidence was uncertain, as many studies administered Pb as
bolus doses. Additionally, few toxicological studies reported the resulting BLLs and those studies that did
provide this evidence had BLLs of limited relevance to environmentally exposed humans (>30 (.ig/dL).
Thus, despite some evidence of Pb-induced hepatic effects, uncertainties related to the relevance of the
available studies limited the causal conclusions that could be drawn in the 2013 Pb ISA.
Recent toxicological studies include more relevant routes of exposure (i.e., drinking water, oral
gavage, and inhalation) and exposures resulting in lower BLLs than those available for the 2013 Pb ISA
(BLL range: 3.6-30.2 (.ig/dL). These studies provide consistent evidence of Pb-induced increases in AST,
ALT, ALP, and GGT activities, which are indicative of reduced liver function (Barkaoui et al.. 2020;
Dumkova et al.. 2020b; Gao et al.. 2020; Andielkovic et al.. 2019; Laamech et al.. 2017; Long et al..
2016; Liu et al.. 2013; Berrahal et al.. 2011). Additionally, recent studies provide consistent evidence of
Pb-associated increases in hepatic oxidative stress, as indicated by decreases in GSH levels and CAT,
SOD, and GPx activities (Barkaoui et al.. 2020; Gao et al.. 2020; Andielkovic et al.. 2019; Laamech et al..
2017; Li et al.. 2017; Long et al.. 2016). While impaired lipid metabolism was reported in the 2013 Pb
ISA, a limited number of recent studies of cholesterol have reported contrasting results, one indicating
Pb-induced increases in total cholesterol (Laamech et al.. 2017) and the other reporting decrements in
total cholesterol (Dumkova et al.. 2020a).
In contrast to toxicological evidence, recent epidemiologic studies evaluating the relationship
between BLLs and hepatic effects are generally inconsistent or inconclusive. Similar to studies evaluated
in the 2013 Pb ISA, most recent studies implement cross-sectional designs, although they include more
robust adjustment for potential confounders and populations with much lower mean BLLs. Still, these
studies do not establish temporality between exposure and outcome or address potentially large
differences in past versus current exposures. There is therefore uncertainty as to the specific timing,
duration, frequency, and level of Pb exposure that contributed to any observed associations. The strongest
evidence for direct liver injury comes from a large cross-sectional analysis of adults in China that reported
a positive association between BLLs and NAFLD prevalence measured by ultrasound (Zhai et al.. 2017).
Other analyses used biomarkers or biomarker indices to assess NAFLD, which are less accurate than
ultrasonic imaging and may introduce non-differential misclassification. Non-differential
misclassification of a dichotomous outcome is likely to bias results toward the null. The available
biomarker studies of NAFLD did not provide convincing evidence that BLLs are associated with NAFLD
prevalence (Chung et al.. 2020; Reia et al.. 2020; Werder et al.. 2020). Results from studies that examined
serum biomarkers of general liver function (e.g., AST, ALT, ALP, GGT, and bilirubin) provided some
evidence that BLLs are positively associated with biomarker levels (Chen et al.. 2019; Obeng-Gvasi.
2019; Pollack et al.. 2015). but the inference that can be drawn from these studies is limited in light of
less consistent evidence from more direct measures of hepatic function. There are also a few recent
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studies that examined serum lipids in adults or children and the results are inconsistent. Across studies,
contrasting associations were observed between BLLs and specific lipids, with no discernable pattern of
associations between BLLs and triglycerides, LDL cholesterol, HDL cholesterol, or total cholesterol.
Overall, the collective evidence is suggestive of, but not sufficient to infer, a causal
relationship between Pb exposure and hepatic effects. This conclusion is based on the strength of the
toxicological evidence and some remaining inconsistencies and uncertainties in the epidemiologic
evidence. Recent toxicological studies build upon evidence from the 2013 Pb ISA and provide largely
consistent evidence that indicates exposure to Pb can result in altered liver function and hepatic oxidative
stress. Compared to the 2013 Pb ISA, recent toxicological studies include routes of exposure and BLLs
that are more relevant to humans. Results from a limited number of recent epidemiologic studies
examining liver enzymes are generally coherent with the toxicological evidence, indicating Pb-associated
increases in enzymes that are consistent with altered liver function. However, due to the reported liver
enzyme levels in the epidemiologic studies and inconsistent evidence of an association between BLLs and
direct liver injury, there is uncertainty as to whether the observed changes in enzymes are indicative of
liver damage. The key evidence, as it relates to the causal framework, is summarized in Table 9-1.
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Table 9-1 Evidence that is suggestive of, but not sufficient to infer, a causal relationship between Pb
exposure and hepatic effects
Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with Effects0
Consistent evidence from Toxicological studies in rodents provide
animal toxicological largely consistent evidence that indicates
studies at relevant BLLs exposure to Pb can result in:
Altered liver function
Increases in hepatic oxidative stress, as
indicated by decreases in GSH levels and
CAT, SOD, and GPx activities
Berrahal etal. (2011)
Liu etal. (2013)
Long etal. (2016)
Andielkovic et al. (2019)
Gao et al. (2020)
Dumkova et al. (2020b)
Laamech et al. (2017)
Barkaoui et al. (2020)
Li etal. (2017)
Long etal. (2016)
Andielkovic et al. (2019)
Barkaoui et al. (2020)
Gao et al. (2020)
Laamech et al. (2017)
Range of mean BLLs across studies:
18.0 to 29.0 |jg/dL
Range of mean BLLs across studies:
3.6 to 30.2 |jg/dL
Limited or inconsistent
evidence from
epidemiologic studies at
relevant BLLs
Inconsistent evidence of associations
between BLLs and NAFLD
Some evidence that BLLs are associated
with increased levels of serum biomarkers
of liver function, but limited inference due to
study populations that had biomarkers well
within normal ranges
See Section 9.1.3.1
Pollack et al. (2015)
Chen etal. (2019)
Obeng-Gvasi (2019)
Range of mean BLLs across studies:
1.0 to 5.29 |jg/dL
Range of mean BLLs across studies:
1.0 to 8.7 |jg/dL
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Rationale for Causality
Determination3
Key Evidence13
References'3
Pb Biomarker Levels Associated with Effects0
Biological Plausibility The proposed pathway leading from Pb See Section 9.1.4
exposure to hepatic health effects begins
with the induction of inflammation and
increase in oxidative stress. This results in
both changes in metabolizing enzymes and
cholesterol synthesis that could be
responsible for fatty accumulation in the
liver. Widespread oxidative damage results
in cell loss which could disrupt the normal
liver structure and contribute to loss of liver
function.
BLLs = blood lead levels; CAT = catalase; GSH = glutathione; GPx = glutathione peroxidase; NAFLD = nonalcoholic fatty liver disease; Pb = lead; SOD = superoxide dismutase.
aBased on aspects considered in judgments of causality and weight-of-evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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9.2
Metabolic Effects
9.2.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current
Review
The 2013 Pb ISA (U.S. EPA. 2013) did not have a separate discussion of potential metabolic
effects of exposure to Pb. However, evidence relevant to metabolic effects was provided by a small
number of studies that examined glucose and insulin homeostasis, lipids, cholesterol, and liver health
endpoints. These studies provided evidence for modes of action and were discussed across a few sections
of the 2013 Pb ISA (U.S. EPA. 2013). including Section 4.4 (Cardiovascular Effects), Section 4.5 (Renal
Effects), and Section 4.9.1 (Effects on the Hepatic System). There was no causality determination for
metabolic effects in the 2013 Pb ISA (U.S. EPA. 2013).
The metabolic effects reviewed in this section include diabetes mellitus and insulin resistance
(Section 9.2.3.1), metabolic syndrome and its components (Section 9.2.3.2), and effects on body weight
measures (Section 9.2.3.3). Other metabolic indicators, such as changes in liver function, serum lipids,
and neuroendocrine signaling, are discussed in other sections of this appendix (Sections 9.2 and 9.4).
9.2.2 Scope
The scope of this section is defined by PECOS statements. The PECOS statement defines the
objectives of the review and establishes study inclusion criteria thereby facilitating identification of the
most relevant literature to inform the Pb ISA.6 In order to identify the most relevant literature, the body of
evidence from the 2013 Pb ISA was considered in the development of the PECOS statements for this
Appendix. Specifically, well-established areas of research; gaps in the literature; and inherent
uncertainties in specific populations, exposure metrics, comparison groups, and study designs identified
in the 2013 Pb ISA inform the scope of this Appendix. The 2013 Pb ISA used different inclusion criteria
than the 2024 Pb ISA, and the studies referenced therein often do not meet the current PECOS criteria
(e.g., due to higher or unreported biomarker levels). Studies included in the 2013 Pb ISA, including many
that do not meet the current PECOS criteria, are discussed in this appendix to establish the state of the
evidence prior to this assessment. Except for supporting evidence used to demonstrate the biological
6The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
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plausibility of Pb-associated metabolic effects, recent studies were only included if they satisfied all of the
components of the following discipline-specific PECOS statements:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect.
Exposure: Exposure to Pb7 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure;8 or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles).
Outcome: Metabolic effects.
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages).
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a BLL of 30 (ig/dL or below.91"
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control.
Outcomes: Metabolic effects.
Study design: Controlled exposure studies of animals in vivo.
7Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area of particular
relevance to the NAAQS review (e.g., longitudinal studies designed to examine recent versus historical Pb
exposure).
8Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 |im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 (im3 (PM2.5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is difficult
to assess the representativeness of these concentrations to population exposure [Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PNLs with BLLs are lacking.
9Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
u'This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLL.
The 95th percentile of the 2011-2016 NHANES distribution of BLL in children (1-5 years; n= 2,321) is 2.66 (ig/dL
(Egan et al„ 2021) and the proportion of individuals with BLL that exceed this concentration varies depending on
factors including (but not limited to) housing age, geographic region, and a child's age, sex, and nutritional status.
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9.2.3
Epidemiologic Studies on Metabolic Effects
9.2.3.1 Diabetes Mellitus and Insulin Resistance
Diabetes mellitus is a chronic condition characterized by an inability to regulate glucose in the
blood by producing or responding to insulin. A number of epidemiologic studies evaluated in the 2013 Pb
ISA (U.S. EPA. 2013) examined diabetes as a potential at-risk factor that could modify the relationship
between Pb exposure and other health outcomes, but none examined the direct relationship between Pb
exposure and diabetes incidence or prevalence. Recent studies have examined this relationship,
commonly categorizing diabetes mellitus status as meeting one or more of the following criteria: (1)
elevated fasting blood glucose (FBG), (2) self-reported use of insulin or oral medications for diabetes, or
(3) self-reported physician diagnosis with diabetes. There are three primary types of diabetes: type I, type
II, and gestational diabetes mellitus (GDM). Some of the evaluated studies distinguished between types of
diabetes mellitus, while others did not. Most studies were cross-sectional in design, meaning temporality
between exposure and outcome could not be established.
Recent epidemiologic studies examining the relationship between Pb exposure and diabetes
mellitus, or insulin resistance have reported mostly null findings across lifestages. In adult populations, a
limited number of case-control and cross-sectional studies examining diabetes prevalence reported null or
inverse associations between BLLs and diabetes mellitus or levels of insulin resistance. Results from
recent studies examining insulin resistance in adolescents and gestational diabetes in pregnant women are
also mostly null. Measures of central tendency for Pb biomarker levels used in each study, along with
other study-specific details, including study population characteristics and select effect estimates, are
highlighted in Table 9-6. An overview of the recent evidence is provided below.
Studies in Adults
In a recent cross-sectional analysis of blood Pb and diabetes using data from the 2009 and 2010
cycles of the KNHANES, Moon (2013) observed a negative trend in diabetes prevalence across blood Pb
quartiles. In reference to the lowest blood Pb quartile (geometric mean (GM): 1.43 |ig/dL). the smallest
Ors were observed in the highest exposure quartile (GM: 4.08 (ig/dL; OR= 0.745 [95% CI: 0.516, 1.077])
and in the second highest quartile (GM: 2.74 (ig/dL) (OR =0.759 [95% CI: 0.531, 1.086]). Similarly, in
stratified analyses examining effect modification by sex in subjects without diabetes, Moon (2013)
reported slightly lower Homeostatic Model Assessment for Insulin Resistance (HOMA-IR), HOMA of |3-
cell function (HOMA-|3), and fasting insulin per log unit higher BLL. The observed results were
comparable in men and in women.
Two recent cross-sectional case-control studies originating from the Nord-Trondelag Health
Study (HUNT3) evaluated differences in blood Pb measurements between subjects with and without type
II diabetes and reported results that are also consistent with a null or negative association (Hansen et al..
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2017; Simic et al.. 2017). Specifically, Hansen et al. (2017) identified 128 cases of previously
undiagnosed, screening-detected type II diabetes and 755 age- and sex-matched controls. The authors
observed slightly higher, but notably imprecise odds of screening-detected type II diabetes for blood Pb
quartile 4 compared to quartile 1 (OR= 1.12 [95% CI: 0.58, 2.16]). As indicated by the wide confidence
intervals, higher odds are difficult to distinguish from chance. In a parallel analysis, Simic et al. (2017)
identified 267 cases of self-reported type II diabetes and 609 frequency-matched controls from the same
HUNT3 cohort. Consistent with results from Moon (2013). (Simic et al.. 2017) observed substantially
lower diabetes prevalence corresponding to participants with BLLs in the highest quartile compared to the
lowest (OR = 0.24 [95% CI: 0.13, 0.47]). The observation of a negative association for Pb and type II
diabetes by Simic et al. (2017) but not Hansen et al. (2017) may be related to differences in exposure
contrast between identified cases and controls. Hansen et al. (2017) reported median BLLs of 1.99 (ig/dL
for controls and 1.94 (ig/dL for cases, while Simic et al. (2017) reported median BLLs of 2.02 (ig/dL for
controls and 1.64 (ig/dL for cases. Additionally, the differences could be due to an effect of diabetes
treatment on BLLs, which highlights an uncertainty of these cross-sectional analyses.
Studies in Adolescents
A recent study assessed the relationship between exposure to Pb in utero and insulin resistance in
adolescence (Liu et al.. 2020). Pregnant mothers were enrolled in the ELEMENT project from 1997-1999
and 2001-2003 and their children were followed until 2015. There was a null association between first
trimester maternal blood Pb >5 (ig/dL and HOMA-IR in adolescence. The authors also examined effect
modification by sex and reported null associations for boys and girls.
Studies in Pregnant Women
A number of recent studies have investigated the relationship between Pb exposure and GDM.
These studies, most of which have reported null associations between BLLs and GDM, are discussed in
more detail in Section 8.4.1.1.2 of the Reproductive and Developmental Effects Appendix.
9.2.3.2 Metabolic Syndrome and its Components
Metabolic syndrome (MetS) describes a set of cardiometabolic conditions that increase a person's
risk for cardiovascular diseases. Components of MetS include elevated blood pressure, low HDL
cholesterol, elevated blood triglycerides, elevated FBG, and a high waist circumference, also referred to
as abdominal obesity. A MetS diagnosis is commonly defined as meeting three or more of the following
criteria: (1) elevated blood pressure (systolic blood pressure >130 mmHg or diastolic blood pressure
>85 mmHg or current use of blood pressure medication); (2) low HDL cholesterol (<40 mg/dL in women
or <50 mg/dL in men); (3) elevated serum triglycerides (>150 mmHg) or current use of anti-dyslipidemia
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medication; (4) elevated FBG (>100 (.ig/dL): (5) abdominal obesity (waist circumference >90 cm in men
or >85 cm in women). None of the studies evaluated in the 2013 Pb ISA (U.S. EPA. 2013) examined the
relationship between Pb exposure and MetS. Recent evidence for the effects of Pb exposure on MetS and
its components is inconsistent. Measures of central tendency for Pb biomarker levels used in each study,
along with other study-specific details, including study population characteristics and select effect
estimates, are highlighted in Table 9-3. An overview of the recent evidence is provided below.
9.2.3.3 Metabolic Syndrome
A number of recent large, population-based cross-sectional studies have analyzed the relationship
between BLLs and MetS prevalence and provide inconsistent evidence of an association. Across studies,
mean and/or median BLLs were below 5 (ig/dL, including some below 2 (ig/dL. Studies analyzing data
from overlapping cycles of the KNHANES observed higher MetS prevalence in participants with higher
BLLs (Moon. 2014; Rhee et al.. 2013). Specifically, Rhee et al. (2013) reported that 2008 KNHANES
participants with BLLS in the highest exposure quartile (3.07-19.43 |ig/L) were 2.57 (95% CI: 1.46, 4.51)
times more likely to have MetS than subjects in the lowest quartile (0.42-1.73 |ig/L). The authors noted a
consistent concentration-response trend across quartiles. In an analysis incorporating more KNHANES
cycles (2007-2012), Moon (2014) observed higher odds of MetS for subjects in the second highest
exposure quartile (GM 2.51 (ig/dL) (OR= 1.21 [95% CI: 0.90, 1.62]) compared to the lowest (GM
1.23 (ig/dL) but did not observe a clear dose-response trend across quartiles.
In contrast to KNHANES studies, other analyses of data from a variety of large population-based
surveys noted negative associations between BLLs and MetS (Wen et al.. 2020; Bulka et al.. 2019; Shim
et al.. 2019). Bulka etal. (2019) used data from two NHANES cycles (2011-2014) to perform a cross-
sectional analysis of blood Pb and MetS prevalence. The authors observed lower odds of MetS at higher
blood Pb quartiles, with the lowest odds observed in subjects in the highest quartile of Pb exposure (1.64-
15.98 (ig/dL) compared to the lowest quartile (0.18-0.70 (ig/dL) (OR = 0.81 [95% CI: 0.64, 1.03]). Shim
et al. (2019) and Wen et al. (2020) similarly reported lower odds of MetS associated with higher BLLs in
the Korean National Environmental Health Survey II (KNHANES II) and a survey of adults in Taiwan,
respectively.
Components of Metabolic Syndrome
In addition to cross-sectional studies evaluating MetS prevalence, several recent studies have
assessed the potential effects of Pb on the individual components of MetS (i.e., abdominal obesity [often
measured by waist circumference], low HDL cholesterol, elevated triglycerides, and elevated FBG;
studies evaluating blood pressure and hypertension are discussed in Section 4.3). Similar to studies that
evaluated MetS prevalence, most of these studies analyzed cross-sectional data from nationally
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representative health surveys. In general, results from recent studies were inconsistent across individual
MetS components, with the exception of blood pressure and serum triglycerides.
Waist Circumference
Recent KNHANES analyses of BLLs and waist circumference were inconsistent (Lee and Kim.
2016. 2013; Rhee et al.. 2013). In an analysis of KNHANES participants from 2005-2010, Lee and Kim
(2013) observed no apparent association between BLLs and waist circumference. The same authors
evaluated more recent KNHANES cycles (2007-2012) and observed slightly higher odds of waist
circumference >85 cm in the second ( >2.199-3.011 |ig/d) and third (>3.011 (ig/dL) blood Pb tertiles
compared to the first tertile (<2.199 (.ig/dL). but slightly lower odds per twofold higher BLLs (Lee and
Kim. 2016). In contrast, in an analysis of 2008 KNHANES participants, Rhee et al. (2013) found a
modest positive association between blood Pb and abdominal circumference as a continuous variable.
Results from two recent NHANES analyses were similarly inconsistent (Bulka et al.. 2019; Wang
et al.. 2018c). Wang et al. (2018c) used data from NHANES cycles between 2003 and 2014 and observed
0.8% (95% CI: 0.6, 1.0%) lower waist circumference per 1-SD higher logurtransformed BLL ((.ig/dL). In
contrast, a study including two NHANES cycles that overlapped with the Wang et al. (2018c) study
(2011-2014) reported negative associations between BLLs and probability of abdominal obesity (Bulka et
al.. 2019).
HDL Cholesterol and Serum Triglycerides
The previously discussed KNHANES analyses also assessed HDL cholesterol and serum
triglycerides. These studies do not provide evidence that BLLs are associated with higher odds of low
HDL cholesterol (Lee and Kim. 2016. 2013; Rhee et al.. 2013). The same studies did provide consistent
evidence of higher serum triglycerides in association with higher BLLs, although these studies were
notably conducted in overlapping populations (i.e., non-independent samples). Lee and Kim (2013) and
Lee and Kim (2016) observed slightly higher odds of high serum triglycerides (>150 (ig/dL) with higher
BLLs (analyzed as a continuous variable and as tertiles). Similarly, Rhee et al. (2013) reported a modest
positive association between serum triglycerides and log-transformed BLLs.
In addition to studies that examined HDL cholesterol and serum triglycerides in conjunction with
MetS, a few other recent studies also evaluated these measures as part of a broader lipids profile. As
discussed in Section 9.1.3.3, these studies were inconsistent for HDL cholesterol and triglycerides,
including a prospective cohort study of older Veterans participating in the NAS that reported null
associations between BLLs at baseline and HDL cholesterol and triglyceride levels after three to four
years of follow-up (Peters et al.. 2012).
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Elevated Fasting Glucose
The majority of recent population-based cross-sectional studies of MetS components did not
observe associations between BLLs and FBG. Specifically, KNHANES analyses (Lee and Kim. 2016;
Rhee et al.. 2013) and a recent NHANES analysis (Bulka et al.. 2019) reported null associations between
BLLs and FBG. In contrast, in an analysis of earlier KNHANES cycles, Lee and Kim (2013) reported
BLLs to be positively associated with elevated FBG (>100 (.ig/dL). with higher odds of elevated FBG
relative to two-fold higher BLLs (OR = 1.118 [95% CI: 0.953, 1.311]). In addition to large cross-sectional
studies, a smaller cross-sectional analysis of adults of African descent across five countries of varying
social and economic development in Africa also examined the relationship between BLLs and elevated
FBG (Ettinger et al.. 2014). Ettinger et al. (2014) reported higher odds of elevated FBG (>100 mg/dL) in
subjects with a blood Pb exposure level above the median (1.66 (ig/dL) compared to those below it
(OR = 4.99 [95% CI: 1.97, 12.69]). However, the small sample size (n = 150) in this study reduces
statistical power and the precision of the effect estimate.
9.2.3.4 Body Weight Measures in Adults
A few epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA. 2013) examined obesity as
a potential risk factor that could modify the relationship between Pb exposure and other health outcomes,
but none examined the direct relationship between Pb exposure and body weight measures in adults.
Recent studies have examined this relationship, commonly assessing body weight using body mass index
(BMI), a measure of body fat that is calculated as a person's weight divided by the square of their height.
For adults, overweight is defined as having a BMI of 25 kg/m2 or greater and obesity is defined as having
a BMI of 30 kg/m2 or greater. Studies examining Pb and body weight measures in children and
adolescents are discussed in the Reproductive and Developmental Effects Appendix of the 2024 Pb ISA
(Section 8.5.1.1).
A limited number of recent studies have examined the relationship between Pb exposure and
body weight measures in adults. Overall, the current evidence for the effects of Pb exposure on body
weight measures is inconsistent, although small sample sizes limit the interpretation of a few of the
studies. Additionally, recent studies are cross-sectional, which reduces confidence in their results because
temporality between exposure and outcome cannot be established. Measures of central tendency for Pb
biomarker levels used in each study, along with other study-specific details, including study population
characteristics and select effect estimates, are highlighted in Table 9-3. An overview of the recent
evidence is provided below.
Recent studies examining Pb exposure and body weight measures in adults utilize cross-sectional
study designs. In an analysis of a large population-based survey of Chinese citizens, Wang et al. (2018a)
observed higher BMI (|3 = 0.24 kg/m2 [95% CI: 0.08, 0.40 kg/m2]) and odds of being overweight or obese
(OR =1.13 [95% CI: 1.02, 1.25]) associated with each natural log unit higher level of blood Pb (|ig/L). In
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order to account for potential reverse causality, the authors used Mendelian randomization to assess the
relationship between BLLs and genetic variants associated with increased BMI. Because the genetic
variants precede exposure, the variants are expected to be associated with BLLs if BMI is a potential
causal factor of increased BLLs. Wang et al. (2018a) reported null associations between BLLs and an
aggregate measure of single nucleotide polymorphisms constructed to represent susceptibility to high
BMIs.
Other recent studies were less informative due to small sample sizes. In a cross-sectional analysis
of adults of African descent across five countries of varying social and economic development in Africa,
Ettinger et al. (2014) compared the prevalence of being overweight (BMI >25) or being obese (BMI >30)
among subjects above versus below the median blood Pb exposure level (1.66 (ig/dL). Among subjects
with above median blood Pb, Ettinger et al. (2014) observed slightly lower odds of being overweight
(OR = 0.88 [95% CI: 0.31, 2.51]), but higher odds of being obese (OR = 2.70 [95% CI: 0.75, 9.75]). The
observed associations, however, were notably imprecise due to the small sample size (n = 150). In
contrast, another small cross-sectional study of 145 adult men living in China observed a null association
between BLLs and BMI (Guo et al.. 2019). As is the case in both of these studies, limited statistical
power resulting from a small sample size reduces statistical power and precision, which might explain the
incongruous results.
9.2.4 Toxicological Studies on Metabolic Effects
The 2013 Pb ISA did not have a section devoted to toxicological studies related to the effect of Pb
on metabolism. However, as discussed in the Section 9.1.4, a few studies evaluated in the 2013 Pb ISA
demonstrated that Pb exposure can impair lipid metabolism in animals, as evidenced by increased hepatic
cholesterogenesis, and altered triglyceride and phospholipid levels (Sharma et al.. 2010; Ademuviwa et
al.. 2009; Khotimchenko and Kolenchenko. 2007). The relevance of the toxicological evidence is
uncertain, as many studies administered Pb as bolus doses and/or results were observed in animals with
high BLLs. In subsequent years, there have been a few PECOS-relevant publications on Pb exposure and
metabolic effects. In general, these studies cover disparate endpoints, but provide some evidence of Pb-
induced changes in metabolic activity in rodents.
In a lifetime study using mice, Faulk et al. (2014) assessed perinatal Pb exposures via Pb acetate
in drinking water from conception to weaning. Average maternal BLLs for exposed groups ranged from
4.1 to 32 (ig/dL. The study findings included sex-specific increases in energy expenditure, food intake,
body weight, total body fat, activity, and insulin response. In addition, a study in weanling rats that
focused on neuropathology found that Pb exposure decreased cholesterol levels in brain tissue (Zhou et
al.. 2018). The latter study, which also used Pb acetate in drinking water, reported BLLs ranging from
14.7 to 28.9 (ig/dL. Finally, in an investigation of the effects of vitamin D metabolism in rats, Rahman et
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al. (2018) reported that Pb interferes with vitamin D metabolism by affecting the expression of its
metabolizing enzymes.
9.2.5 Summary and Causality Determination
There was no causality determination for metabolic effects in the 2013 Pb ISA (U.S. EPA. 2013).
The number of studies examining Pb exposure and metabolic effects has expanded substantially since the
2013 Pb ISA (U.S. EPA. 2013). highlighted by a number of recent epidemiologic studies, as well as a few
animal toxicological studies currently available for review. The focus of this causality determination is on
altered glucose resistance, diabetes mellitus, MetS, and obesity. Notably, there is significant overlap
between components of metabolic health and the cardiovascular and hepatic systems. While blood
pressure and serum lipids are important components of MetS, they are also discussed in detail in the
cardiovascular effects appendix (Appendix 4) and hepatic effects section (Section 9.1), and contribute to
the causality determinations therein. For the metabolic effects causality determination, these endpoints are
considered to the extent that they contribute to a diagnosis of MetS.
There is some evidence from a limited number of animal toxicological studies that exposure to Pb
resulting in BLLs relevant to humans alters cholesterol metabolism (Zhou et al.. 2018) and leads to
increases in body weight, body fat, and insulin response (Faulk et al.. 2014). In contrast, recent
epidemiologic studies are inconsistent across a range of metabolic outcomes and thus not coherent with
the limited toxicological evidence. A limited number of cross-sectional studies examining diabetes
prevalence and insulin resistance in adults reported null (Hansen et al.. 2017; Simic et al.. 2017) and
negative (Moon. 2013) associations with BLLs. Further, results from analyses of MetS in large national
surveys in the United States and Korea were largely inconsistent. Many of these same studies also provide
generally inconsistent evidence of associations between BLLs and individual components of MetS,
though there is substantial epidemiologic and toxicological evidence that exposure to Pb leads to
increased blood pressure and hypertension (Section 4.3). While a limited number of KNHANES analyses
demonstrate consistent associations between BLLs and serum triglycerides (Lee and Kim. 2016. 2013;
Rhee et al.. 2013). these studies include overlapping study populations and therefore do not provide
independent evidence of associations. Additionally, a recent prospective cohort study of older adults
observed null associations between BLLs at baseline and serum triglyceride levels measured three to four
years later (Peters et al.. 2012). Despite observed associations between BLLs and some of the individual
components of MetS, the available evidence examining the cluster of components does not consistently
associate BLLs with MetS. A notable limitation of the current evidence base is the use of concurrent
BLLs as a biomarker for Pb exposure. Given the chronic nature of MetS, a cumulative measure of Pb
exposure, such as bone Pb, might be more relevant.
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Collectively, given the insufficient quantity of toxicological studies and inconsistency in
epidemiologic results, the evidence is inadequate to infer the presence or absence of a causal
relationship between Pb exposure and metabolic effects.
9.3 Effects on the Gastrointestinal System
9.3.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the
Current Review
The 2013 Pb ISA concluded that "because of the insufficient quantity and quality of studies, the
available evidence was inadequate to determine if there is a causal relationship between Pb exposure and
gastrointestinal effects" (U.S. EPA. 2013). There were very few studies evaluated in the 2013 Pb ISA that
examined Pb exposure and gastrointestinal (GI) effects in humans or animals. Epidemiologic evidence of
an association between Pb exposure and GI effects was limited to a small number of occupational studies
of prevalent symptoms in Pb-exposed workers. The internal validity and generalizability of these studies
was limited by cross-sectional study designs, lack of consideration for potential confounders, and notably
higher BLLs (>40 (ig/dL) than those experienced by the general population. In addition to the
epidemiologic evidence, there were a limited number of toxicological studies that provide evidence of Pb-
induced effects on mechanisms underlying GI damage and impaired function.
9.3.2 Scope
The scope of this section is defined by PECOS statements. The PECOS statement defines the
objectives of the review and establishes study inclusion criteria thereby facilitating identification of the
most relevant literature to inform the Pb ISA.11 In order to identify the most relevant literature, the body
of evidence from the 2013 Pb ISA was considered in the development of the PECOS statements for this
Appendix. Specifically, well-established areas of research; gaps in the literature; and inherent
uncertainties in specific populations, exposure metrics, comparison groups, and study designs identified
in the 2013 Pb ISA inform the scope of this Appendix. The 2013 Pb ISA used different inclusion criteria
than the 2024 Pb ISA, and the studies referenced therein often do not meet the current PECOS criteria
(e.g., due to higher or unreported biomarker levels). Studies included in the 2013 Pb ISA, including many
that do not meet the current PECOS criteria, are discussed in this appendix to establish the state of the
11 The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
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evidence prior to this assessment. Except for supporting evidence used to demonstrate the biological
plausibility of Pb-associated effects on the gastrointestinal system, recent studies were only included if
they satisfied all of the components of the following discipline-specific PECOS statements:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect.
Exposure: Exposure to Pb12 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure13; or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles).
Outcome: Effects on the gastrointestinal system.
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials, and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (i.e., mouse, rat, Guinea pig,
minipig, rabbit, cat, dog; whole organism) of any lifestage (including preconception, in utero,
lactation, peripubertal, and adult stages).
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a BLL of 30 (ig/dL or below.1415
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control.
Outcomes: Effects on the gastrointestinal system.
12Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area of particular
relevance to the NAAQS review (e.g., longitudinal studies designed to examine recent versus historical Pb
exposure).
13Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 |im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 |im3 (PNL 5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is difficult
to assess the representativeness of these concentrations to population exposure [Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PNLs with BLLs are lacking.
14Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
15This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLL.
The 95th percentile of the 2011-2016 NHANES distribution of BLL in children (1-5 years; n= 2,321) is 2.66 (ig/dL
(Egan et al„ 2021) and the proportion of individuals with BLL that exceed this concentration varies depending on
factors including (but not limited to) housing age, geographic region, and a child's age, sex, and nutritional status.
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Study design: Controlled exposure studies of animals in vivo.
9.3.3 Epidemiologic Studies on the Gastrointestinal System
The epidemiologic evidence evaluated in the 2013 Pb ISA was limited to a small number of
occupational cohort studies of prevalent GI symptoms in Pb-exposed workers (U.S. EPA. 2013). As noted
in Section 9.3.1, these studies had a number of limitations, including cross-sectional study designs, lack
of consideration for potential confounders, and notably higher BLLs (>40 (ig/dL) than those experienced
by the general population. There are no recent PECOS-relevant epidemiologic studies that evaluate
potential associations between exposure to Pb and effects on the gastrointestinal system. A limited
number of studies reported associations between BLLs and gut microbiota diversity, as discussed in the
Immune System Effects Appendix (Section 6.6). However, these studies do not inform the relationship
between Pb exposure and specific GI health effects.
9.3.4 Toxicological Studies on the Gastrointestinal System
In the 2013 Pb ISA (U.S. EPA. 2013). specific attention was drawn to a pair of rat studies; one
reporting frequency-dependent inhibition of electric field-stimulated relaxations to nonadrenergic
noncholinergic (NANC) nerve stimulation in rat gastric fundus (possibly due to the modulated release of
NO), and the other focusing on Pb-induced oxidative stress in the gastric mucosa, wherein an increase in
gastric mucosal damage induced by the acidified ethanol was observed. Neither of these studies reported
BLLs. Neither of the two pertinent studies since the 2013 Pb ISA directly addresses these findings
ITReddv et al.. 2018; Kosik-Bogacka et al.. 2011); see below].
In a chronic exposure study with rats, Kosik-Bogacka et al. (2011) confirmed an inhibitory effect
of Pb on electrophysiological parameters, among other findings. These findings were strengthened by
results showing the ability of L-ascorbic acid to (at least partially) abrogate the effects of Pb exposure.
Mean BLLs in this study were reported at 7 (ig/dL.
In a 2018 microbiome study, Reddv et al. (2018) found that Pb-exposed rats had decreased 8-
aminolevulinic acid dehydratase (ALAD) activity and intestinal lactobacillus levels, irrespective of the
dietary iron supplementation. Withdrawal of Pb exposure increased lactobacilli, whereas re-exposure to
Pb decreased lactobacilli population. BLLs were reported in the range of 19 to 48 (ig/dL.
9.3.5 Summary and Causality Determination
The 2013 Pb ISA concluded that evidence was "inadequate" to determine whether a causal
relationship exists between Pb exposure and GI effects (U.S. EPA. 2013). This causality determination
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was based on an insufficient quantity and quality of studies in the cumulative body of evidence. A limited
number of occupational cohort studies indicated associations between BLLs and prevalent symptoms,
such as stomach pain, gastritis, constipation, and intestinal paralysis. However, the implications of these
findings are limited by the cross-sectional study designs, high BLLs associated with effects (mostly
>40 (ig/dL), and limited consideration of potential confounding by factors such as age, smoking, alcohol
use, nutrition, or other occupational exposures. Toxicological evidence indicates that Pb is absorbed
primarily in the duodenum by active transport and diffusion, although variability is observed by Pb
compound, age of intake, and nutritional factors. There was some coherence between the evidence in Pb-
exposed workers and observations in animals that Pb induces damage to the intestinal mucosal
epithelium, decreases duodenum contractility and motility, reduces absorption of calcium ion(s) (Ca2+),
inhibits NANC relaxations in the gastric fundus, and induces oxidative stress (lipid peroxidation,
decreased SOD and CAT) in the gastric mucosa.
Recent studies are limited in number, and while some provide potential biological plausibility for
Pb-induced GI effects, none directly inform the relationship between Pb exposure and GI effects. Given
the insufficient quantity and quality of studies, the evidence remains inadequate to infer the presence or
absence of a causal relationship between Pb exposure and gastrointestinal effects.
9.4 Effects on the Endocrine System
9.4.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current
Review
The 2013 Pb ISA (U.S. EPA. 2013) evaluated a limited number of studies examining the
relationship between exposure to Pb and effects on the endocrine system. Epidemiologic and
toxicological evidence related to male and female sex hormones, which was generally inconsistent, is
discussed in more detail in Appendix 8 (Sections 8.6.1.1 and 8.7.2). In addition to studies on sex
hormones, results from a small number of epidemiologic and toxicological studies on Pb-associated
endocrine effects such as changes in thyroid hormones, Cortisol, corticosterone, and vitamin D levels were
also inconsistent. Further, epidemiologic studies were mostly cross-sectional and included limited
consideration for potential confounders. As a whole, the limited quantity, quality, and consistency of the
available evidence was "inadequate to determine if there is a causal relationship between Pb exposure and
endocrine effects related to thyroid hormones, Cortisol, and vitamin D.
9.4.2 Scope
The scope of this section is defined by PECOS statements. The PECOS statement defines the
objectives of the review and establishes study inclusion criteria thereby facilitating identification of the
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most relevant literature to inform the Pb ISA.16 In order to identify the most relevant literature, the body
of evidence from the 2013 Pb ISA was considered in the development of the PECOS statements for this
Appendix. Specifically, well-established areas of research; gaps in the literature; and inherent
uncertainties in specific populations, exposure metrics, comparison groups, and study designs identified
in the 2013 Pb ISA inform the scope of this Appendix. The 2013 Pb ISA used different inclusion criteria
than the 2024 Pb ISA, and the studies referenced therein often do not meet the current PECOS criteria
(e.g., due to higher or unreported biomarker levels). Studies included in the 2013 Pb ISA, including many
that do not meet the current PECOS criteria, are discussed in this appendix to establish the state of the
evidence prior to this assessment. Except for supporting evidence used to demonstrate the biological
plausibility of Pb-associated effects on the gastrointestinal system, recent studies were only included if
they satisfied all of the components of the following discipline-specific PECOS statements:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect.
Exposure: Exposure to Pb17 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure18; or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles).
Outcome: Effects on the endocrine system.
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
16The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
17Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area of particular
relevance to the NAAQS review (e.g., longitudinal studies designed to examine recent versus historical Pb
exposure).
18Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 |im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 (im3 (PM2.5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is difficult
to assess the representativeness of these concentrations to population exposure [Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PNfc.s with BLLs are lacking.
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Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages).
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal {in vivo) that
results in a BLL of 30 (ig/dL or below.19'2"
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control.
Outcomes: Effects on the endocrine system.
Study design: Controlled exposure studies of animals in vivo.
9.4.3 Epidemiologic Studies on the Endocrine System
A limited number of epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA. 2013)
reported associations between exposure to Pb and measures of endocrine function, including thyroid
hormones, Cortisol, and vitamin D levels. However, most studies were cross-sectional in design, and
many did not consider potential confounding factors. Further, while some studies did find associations
between Pb exposure and endocrine effects, the results for specific hormones were not consistent.
There are several recent epidemiologic studies of Pb exposure and endocrine function, which also
implement cross-sectional analyses but included more robust adjustment for potential confounding
factors. The majority of recent studies are large NHANES analyses that provide generally consistent
evidence of null associations between Pb exposure and thyroid hormone and Cortisol levels. However,
given that these studies examined overlapping study populations, the generally consistent results across
these studies should not be considered independent evidence of a null association. Most recent studies
evaluated potential associations between Pb exposure and thyroid hormone levels, including
triiodothyronine (T3), thyroxine (T4), and thyroid-stimulating hormone (TSH). There were a few studies
that looked at associations between Pb exposure and Cortisol levels and no recent PECOS-relevant studies
that looked at Pb exposure and vitamin D levels. Measures of study-specific BLLs and endocrine effect
estimates are highlighted in Table 9-9. An overview of recent evidence is provided below.
The most consistent evidence from recent studies indicates null associations between BLLs and
TSH, T3, and free T4 (FT4) levels in adults. A few recent NHANES analyses, which included nationally
representative study populations of adults over 20 years old, reported null associations between BLL and
19Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
2"This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLL.
The 95th percentile of the 2011-2016 NHANES distribution of BLL in children (1-5 years; n= 2,321) is 2.66 (ig/dL
(Egan et al.. 2021) and the proportion of individuals with BLL that exceed this concentration varies depending on
factors including (but not limited to) housing age, geographic region, and a child's age, sex, and nutritional status.
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TSH levels in adults (Krieg. 2019; Chen et al.. 2013; Mendv et al.. 2013; Christensen. 2012). Recent
NHANES analyses also provide generally consistent evidence of null associations between BLLs and
FT4 levels (Luo and Hendrvx. 2014; Chen et al.. 2013; Mendv et al.. 2013) as well as between blood Pb
and T3 levels (Nie et al.. 2017; Luo and Hendrvx. 2014; Chen et al.. 2013; Mendv et al.. 2013;
Christensen. 2012).
Recent NHANES studies evaluating a potential association between BLLs and total T4 levels
were less consistent. While some recent studies reported null associations between BLLs and T4 levels in
adults (Luo and Hendrvx. 2014; Chen et al.. 2013). others observed negative associations (Krieg. 2019;
Mendv et al.. 2013; Christensen. 2012). For example, Mendv et al. (2013) reported that T4 levels were
0.162 (ig/dL lower (95% CI: -0.321, -0.004 (ig/dL) for each 1 (ig/dL higher level of blood Pb.
Additionally, while Luo and Hendrvx (2014) noted a null association between BLLs and T4 levels in the
total population, the authors observed a significant negative association between blood Pb and T4 levels
among men after stratifying by sex. Krieg (2019) also found a negative association between blood Pb and
T4 levels, reporting 38.91% (95% CI: -51.25, -23.44) lower T4 levels for each 1 (ig/dL higher level of
blood Pb.
A limited number of NHANES analyses evaluated potential associations between blood Pb and
free T3 (FT3) levels (Luo and Hendrvx. 2014; Chen et al.. 2013; Mendv et al.. 2013). In an analysis of
adults, Mendv et al. (2013) reported a null association between blood Pb and FT3 levels in the general
adult population. This is consistent with the findings of Chen et al. (2013). who reported a null
association between BLLs and FT3 levels in both adolescents (12-19 years old) and adults (>20 years
old). Both studies performed analyses on the 2007-2008 continuous NHANES cycle. Luo and Hendrvx
(2014) evaluated 2007-2010 data, reporting a positive association between blood Pb and FT3 in the
general adult population. The authors reported that FT3 levels were 0.04 (ig/dL (95% CI: 0.01, 0.08)
higher in the highest tertile of blood Pb compared to the lowest.
In addition to the NHANES analyses discussed above, another recent cross-sectional study
examined the relationship between BLLs and thyroid hormone levels in a small study of pregnant women
(n = 291) from the Yugoslavia Prospective Study of Environmental Lead Exposure Cohort (Kahn et al..
2014). Kahn et al. (2014) reported a null association between BLL and TSH levels and a negative
association between blood Pb and FT4 levels.
Two recent cross-sectional studies examined associations between BLLs and Cortisol levels
(Ngueta et al.. 2018; Souza-Talarico et al.. 2017). In a small study of older adults (n = 65) in Montreal,
Canada, Ngueta et al. (2018) reported null associations between BLLs and both diurnal and stress-reactive
Cortisol secretion. In contrast, another small study of non-occupationally exposed Brazilian older adults
(n = 126), Souza-Talarico et al. (2017) reported positive associations between BLLs and both Cortisol
awakening response (CAR) and overall Cortisol concentration. The authors reported that CAR was
0.791 (ig/dL (95% CI: 0.672, 1.073 (ig/dL) higher per 1 (ig/dL higher level of blood Pb. However, it is
worth noting that participants showed an elevated basal circadian level of salivary Cortisol independent of
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Pb exposure, suggesting this population has more repeated exposure to stressful events. Furthermore,
while all participants were older postmenopausal adults, sex was unevenly represented with n = 105
(83%) of the participants being women.
9.4.4 Toxicological Studies on the Endocrine System
The 2013 Pb ISA summarized a few toxicological studies that reported on effects of Pb exposure
on the endocrine system. Specifically, T3 and T4 levels were found to be elevated in cows that were
grazing on land near Pb/operational Zn smelters when compared with cows grazing in unpolluted areas
(Swarup et al.. 2007). However, when regression analyses were done to evaluate potential associations
between BLLs and plasma Cortisol levels in these same cows, no association was observed. Another study
conducted in Wistar rats reported that 21 days of intraperitoneal (i.p.) injections with 8.0 mg/kg Pb led to
increased corticosterone levels and adrenal weights [BLLs not reported; (Biswas and Ghosh. 2006)1.
Some recent studies have also investigated the effects of Pb on the endocrine system (Table 9-5). The
only studies that investigated adrenal gland weight were conducted in Sprague Dawley rats that were
dosed from postnatal day (PND) 4 to 28 and reported no effect of Pb treatment on the weight of the
adrenal glands [BLLs 3.27-12.5 (ig/dL; (Amos-Kroohs et al.. 2016; Graham et al.. 2011)1. Findings
concerning corticosterone levels in recent studies are equivocal. Some studies reported increased
corticosterone in rats exposed to Pb. Specifically, one study that dosed Long-Evans rats starting prior to
conception until 304 days of age reported increases in corticosterone levels in female rats at 2 months of
age but reported no changes in males at any time point [BLLs 11.3 (ig/dL on PND 61 in females; (Rossi-
George et al.. 2011)1. Another study measured corticosterone levels in Sprague Dawley rats at different
intervals following a shallow water stressor. This study reported that treatment with Pb from PND 4 to 28
increased corticosterone levels in male and female rats 0, 30, and 60 minutes following the stressor on
PND 11,0 and 30 minutes following the stressor on PND 19, and 0 and 30 minutes following the stressor
on PND 29 [BLLs 3.2-12.5 (ig/dL on PND 29; (Graham et al.. 2011)1. A single study reported decreases
in corticosterone in F3 female C57 BL/6 mice whose F1 sires were exposed to Pb from gestational day
(GD) -61 to PND 21 [BLLs 0.4 (ig/dL on PND 6-7; (Sobolewski et al.. 2020)1. Contrasting these studies
are those that did not report any effects of Pb exposure on corticosterone levels. Interestingly, these
studies used similar dosing paradigms to those that reported effects with one study dosing C57 BL/6 mice
starting preconceptionally through adulthood [ending on PND 365; (Corv-Slechta et al.. 2013)1 and the
other study dosing Sprague Dawley rats from PND 4 to 28 (Amos-Kroohs et al.. 2016). and neither study
reported alterations of corticosterone levels in either sex.
9.4.5 Summary and Causality Determination
The 2013 Pb ISA concluded that the evidence was inadequate to determine if there is a causal
relationship between Pb exposure and endocrine effects related to changes in levels of thyroid hormones,
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cortisol/corticosterone, and vitamin D. This causality determination was based on an insufficient quantity
and quality of studies that provided inconsistent or inconclusive evidence for Pb-related endocrine effects.
Epidemiologic evidence presented in the 2013 Pb ISA regarding the effects of Pb on Cortisol levels
consisted of a single study showing a positive association between prenatal Pb exposure and salivary
Cortisol levels in children following an acute stressor (Gump et al.. 2008). The few epidemiologic studies
investigating associations between Pb and thyroid hormone levels presented in the 2013 Pb ISA reported
inconsistent associations. Toxicological evidence in the 2013 Pb ISA regarding the effects of Pb on the
endocrine system in animals was sparse. Biswas and Ghosh (2006) reported that Pb exposure increased
corticosterone levels and adrenal gland weights in Wistar rats. A single study evaluating thyroid hormone
levels in animals summarized in the 2013 Pb ISA reported no clear associations between Pb exposure and
thyroid hormone levels in cattle with environmental exposure to Pb (Swarup et al.. 2007).
Recent epidemiologic and toxicological evidence evaluating the effects of Pb exposure on the
endocrine system continues to be limited and inconsistent. Most recent epidemiologic studies measured
associations between BLLs and thyroid hormone levels. Results from these studies were mostly null,
though there was some evidence of an inverse association between BLLs and T4 levels in a few studies
(Krieg. 2019; Mendv et al.. 2013; Christensen. 2012). and a single study noted sex-specific associations
between BLLs and T4 and FT4 levels (Luo and Hendrvx. 2014). Although most studies reported null
associations, the analyses included overlapping study populations, so they should not be interpreted as
independent evidence of a null association. Additionally, consistent with the studies evaluated in the 2013
Pb ISA, recent studies are cross-sectional in design, which introduces uncertainty regarding the
temporality between exposure and outcome. No recent toxicological studies investigating the effects of Pb
on thyroid hormone levels were available. Only a few recent epidemiologic studies examined Pb effects
on Cortisol levels (Ngueta et al.. 2018; Souza-Talarico et al.. 2017). Results from these studies were
inconsistent, and both had small sample sizes. Multiple toxicological studies reported on the effects of Pb
exposure on corticosterone levels in animals, but results are equivocal. One study reported decreases
(Sobolewski et al.. 2020). two studies reported increases (Graham et al.. 2011; Rossi-George et al.. 2011).
and two studies reported no effect (Amos-Kroohs et al.. 2016; Corv-Slechta et al.. 2013) on corticosterone
levels in Pb-intoxicated animals. In terms of the effects of Pb on adrenal gland weights in animals, only
two recent studies investigated the effects of Pb on adrenal gland weight. These studies reported no
effects of Pb on adrenal gland weight in Sprague Dawley rats (Amos-Kroohs et al.. 2016; Graham et al..
2011). contrasting with the only study that investigated adrenal gland weights in the 2013 Pb ISA (Biswas
and Ghosh. 2006) which reported increased adrenal gland weights. This contrast may be due to variability
in route of exposure used in the experimental design leading to differences in BLLs between the animals
in Biswas and Ghosh (2006). and the more recent studies. Specifically, Biswas and Ghosh (2006) dosed
animals with 8 mg/kg/d of Pb via i.p. injection, whereas the most recent publications dosed animals with
either 1 or 10 mg/kg/d of Pb b via oral gavage (Amos-Kroohs et al.. 2016) or indirectly dosed animals via
Pb in the milk from their dams which were dosed via oral gavage (Graham et al.. 2011). No recent
PECOS-relevant epidemiologic or toxicological studies were identified that measured vitamin D levels.
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In conclusion, recent epidemiologic and toxicological studies continue to provide limited and
inconsistent evidence for endocrine system effects associated with Pb exposure. Due to the insufficient
quantity and quality of the studies available for review and the inconsistent results across those studies,
the evidence remains inadequate to infer the presence or absence of a causal relationship between Pb
exposure and endocrine system effects, including changes in thyroid hormones, cortisol/corticosterone,
and vitamin D levels.
9.5 Effects on the Musculoskeletal System
9.5.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current
Review
The 2013 Pb ISA evaluated the effects of Pb exposure on bone and teeth (U.S. EPA. 2013). In
order to be more inclusive of other health effects related to bone and teeth, the 2024 Pb ISA expands the
considered health outcomes to include effects on the entire musculoskeletal system. The musculoskeletal
system consists of the bones, teeth, muscles, joints, cartilage, and other connective tissues that support the
body, allow for movement, and protect vital organs. Primary effects on the musculoskeletal system
include increases in osteoporosis, increased frequencies of falls and fractures, changes in bone cell
function as a result of replacement of bone calcium with Pb, and depression in early bone growth. Other
effects include tooth loss and periodontitis. Mechanistic evidence from toxicological studies includes
effects on cell proliferation, procollagen type I production, intracellular protein, and osteocalcin in human
dental pulp cell cultures.
A small body of epidemiologic studies evaluated in the 2013 Pb ISA (U.S. EPA. 2013) provided
consistent evidence of associations between Pb biomarker levels and various effects on bone and teeth,
including an increase in osteoporosis, increased frequencies of falls and fractures, tooth loss, and
periodontitis. The results from these studies, adjusting for potential confounding by age and SES-related
factors, were supported by strong toxicological evidence evaluated in the 2013 Pb ISA and the 2006 Pb
AQCD (U.S. EPA. 2006). which reported effects in bone and teeth in animals following Pb exposure.
Exposure of animals to Pb during gestation and the immediate postnatal period was reported to
significantly depress early bone growth with the effects showing concentration-dependent trends.
Systemic effects of Pb exposure included disruption in bone mineralization during growth, alteration in
bone cell differentiation and function due to alterations in plasma levels of growth hormones and
calcitropic hormones such as l,25-[OH]2D3 and impact on Ca2+- binding proteins and increases in Ca2+
and phosphorus concentrations in the bloodstream. Bone cell cultures exposed to Pb had altered vitamin
D-stimulated production of osteocalcin accompanied by inhibited secretion of bone-related proteins such
as osteonectin and collagen. In addition, Pb exposure caused suppression in bone cell proliferation most
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likely due to interference from factors such as growth hormone (GH), epidermal growth factor (EGF),
transforming growth factor-beta 1 (TGF-|31), and parathyroid hormone-related protein (PTHrP).
As in bone, Pb exposure was found to easily substitute for Ca2+ in the teeth and was taken up and
incorporated into developing teeth in experimental animals. Since teeth do not undergo remodeling like
bones do during growth, most of the Pb in the teeth remains in a state of permanent storage. Pb has also
been shown to decrease cell proliferation, procollagen type I production, intracellular protein, and
osteocalcin in human dental pulp cell cultures. Adult rats exposed to Pb have exhibited an inhibition of
the posteruptive enamel proteinases, delayed teeth eruption times, as well as a decrease in microhardness
of surface enamel. Further discussion of these processes and effects, including corresponding references,
can be found in sections 5.8.7 through 5.8.13 of the 2006 Pb AQCD (U.S. EPA, 2006).
In considering the weight of the evidence, the 2013 Pb ISA (U.S. EPA, 2013) concluded that "a
causal relationship is likely to exist between Pb exposure and effects on bone and teeth."
9.5.2 Scope
The scope of this section is defined by PECOS statements. The PECOS statement defines the
objectives of the review and establishes study inclusion criteria, thereby facilitating identification of the
most relevant literature to inform the Pb ISA.21 In order to identify the most relevant literature, the body
of evidence from the 2013 Pb ISA was considered in the development of the PECOS statements for this
Appendix. Specifically, well-established areas of research; gaps in the literature; and inherent
uncertainties in specific populations, exposure metrics, comparison groups, and study designs identified
in the 2013 Pb ISA inform the scope of this Appendix. The 2013 Pb ISA used different inclusion criteria
than the 2024 Pb ISA, and the studies referenced therein often do not meet the current PECOS criteria
(e.g., due to higher or unreported biomarker levels). Studies included in the 2013 Pb ISA, including many
that do not meet the current PECOS criteria, are discussed in this appendix to establish the state of the
evidence prior to this assessment. Except for supporting evidence used to demonstrate the biological
plausibility of Pb-associated effects on the musculoskeletal system, recent studies were only included if
they satisfied all of the components of the following discipline-specific PECOS statements:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect.
21The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
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Exposure: Exposure to Pb22 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure23; or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles).
Outcome: Effects on the musculoskeletal system.
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages).
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a BLL of 30 (ig/dL or below.24,25
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control.
Outcomes: Effects on the musculoskeletal system.
Study design: Controlled exposure studies of animals in vivo.
9.5.3 Epidemiologic Studies on the Musculoskeletal System
A limited number of cross-sectional epidemiologic studies evaluated in the 2013 Pb ISA (U.S.
EPA. 2013) provided consistent evidence of associations between Pb biomarker levels and osteoporosis
and tooth loss after adjusting for potential confounding by age and SES-related factors. Uncertainties in
—Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area of particular
relevance to the NAAQS review (e.g., longitudinal studies designed to examine recent versus historical Pb
exposure).
23Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 |im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 |im3 (PM2.5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is difficult
to assess the representativeness of these concentrations to population exposure [Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PNLs with BLLs are lacking.
24Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
25This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLL.
The 95th percentile of the 2011-2016 NHANES distribution of BLL in children (1-5 years; n= 2,321) is 2.66 (ig/dL
(Egan et al„ 2021) and the proportion of individuals with BLL that exceed this concentration varies depending on
factors including (but not limited to) housing age, geographic region, and a child's age, sex, and nutritional status.
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the evidence base included limited consideration of potential confounding by nutritional factors, a lack of
temporality between exposure and outcome, and uncertainty in the level, timing, frequency, and duration
of Pb exposure that contributed to the observed associations. Recent epidemiologic studies of the
musculoskeletal system generally examine one of three groups of endpoints: (1) bone mineral density
(BMD); (2) joint degeneration; and (3) oral health. Results from recent studies, which adjust for a range
of potential confounders, provide generally consistent evidence of an association between BLLs and
osteoporosis, osteoarthritis, dental caries, and periodontal disease. Recent studies evaluating
musculoskeletal effects are largely cross-sectional analyses, which are unable to establish temporality
between exposure and outcome. Additionally, with biomarkers of Pb exposure, it is difficult to
characterize the specific timing, duration, frequency, and level of Pb exposure that contributed to the
observed associations. This uncertainty may apply particularly to assessments of BLLs, which in
nonoccupationally-exposed adults, reflect both current exposures and cumulative Pb stores in bone that
are mobilized during bone remodeling. Measures of central tendency for Pb biomarker levels used in each
study, along with other study-specific details, including study population characteristics and select effect
estimates, are highlighted in Table 9-11. An overview of the recent evidence is provided below.
9.5.3.1 Bone Mineral Density
A number of recent cross-sectional studies provide generally consistent evidence of an
association between exposure to Pb and BMD in adults. In these studies, BMD (g/cm2) was measured via
X-ray absorptiometry or ultrasound and often converted to a standardized score (i.e., z- and Z-scores)26.
Osteoporosis and osteopenia are characterized by varying degrees of BMD decrements that can
compromise bone microarchitecture. A z-score below -1 often corresponds to osteopenia, whereas a z-
score below -2.0 to -2.5 is categorized as osteoporosis. There are significant sex and age differences in
the incidence of osteoporosis and osteopenia, with postmenopausal women being at greatest risk for
declines in BMD. Because osteoporosis and osteopenia are more common in women, many of the recent
epidemiologic studies evaluating the relationship between BLLs and BMD are either conducted in study
populations comprised of older women or stratified by sex. Importantly, the cross-sectional nature of the
studies does not rule out the possibility that the association is driven by increased BLLs due to higher
bone turnover in individuals with osteoporosis. Additionally, although most analyses include study
populations with mean BLLs <3 (ig/dL, study participants were born prior to the phase-out of leaded
gasoline and therefore likely had much higher past Pb exposures, making it difficult to characterize the
specific timing, duration, frequency, and level of Pb exposure that contributed to the observed
associations.
26Standardized scores are used to analyze BMD data as deviations from average BMD in matched healthy
populations. Underlying populations vary by study.
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A few recent analyses of data from large, nationally representative health surveys provide
generally consistent evidence of an association between BLLs and BMD in women (Wang et al.. 2019;
Cho et al.. 2012; Lee and Kim. 2012). In an analysis of 2008 KNHANES data, Cho et al. (2012) observed
higher odds of osteoporosis associated with higher BLL quartiles in postmenopausal women. The authors
noted associations at low levels (e.g., Q2 [1.83 to <2.32 |ig/dL| versus quartile 1 [<1.83 |ig/dL|) that were
similar in magnitude to comparisons between the higher quartiles and the first quartile, suggesting a
potentially non-linear association. In a similar study, Lee and Kim (2012) analyzed data from the same
KNHANES cycle but expanded the age range to include premenopausal women. The authors reported
that higher BLLs were associated with lower BMD at several bone sites. Additionally, Pb-related BMD
decrements were consistently higher in postmenopausal women compared to premenopausal women. For
example, 1 (ig/dL higher levels of blood Pb were associated with 0.28 g/cm2 (95% CI: 0.11, 0.45 g/cm2)
lower femoral BMD in postmenopausal women compared to 0.15 g/cm2 (95% CI: -0.03, 0.33 g/cm2)
lower femoral BMD in premenopausal women.
In contrast to KNHANES analyses, an analysis of more recent NHANES cycles (2013-2014)
observed null associations between BLLs and BMD in postmenopausal women (Wang et al.. 2019).
Notably, the authors did not control for hormone therapy, which prevents BMD loss and could impact
BLLs due to changes in bone turnover rates. Wang et al. (2019) did note that a higher BLLs were
associated with slightly lower femoral (-0.06 g/cm2 [95% CI: -0.08, -0.03 g/cm2]) and spinal
(-0.05 g/cm2 [95% CI: -0.08, -0.02 g/cm2]) BMD in premenopausal women, as well as higher 10-year
fracture risk scores in the total population (including adult men and women). The findings in
premenopausal women are somewhat consistent with a recent cross-sectional analysis of premenopausal
women in western New York that observed 0.02 (-0.02, 0.05) g/cm2 lower spinal BMD associated with
1 (ig/dL higher BLLs (Pollack et al.. 2013). However, in contrast to the results from Wang et al. (2019).
Pollack et al. (2013) reported null associations between BLLs and total hip and wrist BMD in
premenopausal women.
In a smaller cross-sectional analysis of adults from two communities in southwestern China,
including one with a history of Pb mining and smelting, Li et al. (2020b) observed some evidence of sex-
specific differences in Pb-associated BMD levels. Specifically, female study participants with BLLs
>3.4 (ig/dL had higher odds of osteoporosis compared to female study participants with BLLs <3.4 (ig/dL
(OR= 1.33 [95% CI: 0.61, 2.88]); whereas an inverse association was reported for men (OR= 0.60 [95%
CI: 0.24, 1.49]). However, given the imprecise effect estimates (i.e., wide 95% els), it is difficult to draw
firm conclusions on these sex-specific comparisons.
Other recent studies evaluated the relationship between Pb exposure and BMD in analyses
combining men and women. The inferences that can be drawn from these studies are limited due to
established sex-specific differences in osteoporosis incidence. In an analysis of 2008-2011 KNHANES
cycles, Lim et al. (2016) observed higher odds of osteoporosis or osteopenia across higher BLL quartiles,
with the highest odds noted in quartile 4 (>2.93 (ig/dL) compared to quartile 1 (<1.66 (ig/dL; OR = 1.49
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[95% CI: 1.12, 1.98]). In amuch smaller study ofKorean adults, Lee and Park (2018) similarly reported a
negative association between BLLs and BMD t-scores that was greater in magnitude in participants with a
history of smoking (-0.472 [95% CI-0.85, -0.094]) compared to non-smokers (-0.148 [95% CI: -0.369,
0.073]). The authors also examined over 344,396 single nucleotide polymorphisms (SNPs) mapped to
gene-coding regions to assess potential effect modification of the relationship between BLLs and BMD.
Observed interactions between BLLs and genetic variations were inconsistent after adjustment for
multiple testing, but many implicated interactions with genes and pathways involved in angiogenesis,
bone mass, and nuclear receptor signaling, providing areas of interest for exploring possible mechanisms
that may underlie the observed relationship between BLLs and osteoporosis.
9.5.3.2 Osteoarthritis
A few recent cross-sectional studies examined the association between BLLs and osteoarthritis
(OA) in adults. In an analysis of multiple KNHANES cycles (2010-2012), Park and Choi (2019) reported
that higher natural log BLLs were associated with higher odds of radiographic and symptomatic knee OA
(radiographic osteoarthritis [rOA] and symptomatic osteoarthritis [sxOA]) in postmenopausal women
(OR= 1.77 [95% CI 1.17, 2.67] and 1.50 [95% CI: 0.90, 2.53], respectively). There is some evidence that
the association is mediated by BMI, but there is evidence of a direct association as well (i.e., adjusted for
BMI). The authors noted null associations between BLLs and back OA.
In a cross-sectional analysis of African American and white adults, Nelson et al. (2011b) also
observed associations between BLL and rOA and sxOA in the knee. In a similar study, the same group
noted associations between BLLs and some biomarkers of joint tissue metabolism, including NTX-I,
which is responsible for bone turnover; CTX-II, which is associated with prevalence of rOA in the knee;
COMP (cartilage oligomeric matrix protein), which is a cartilage biomarker; and CPU (carboxypropeptide
of type II collagen), which is linked with collagen synthesis (Nelson et al.. 201 la). Notably, the authors
examined a wide range of biomarkers and stratified their models by sex to examine potential effect
modification, increasing the likelihood of multiple testing bias.
Although all of the studies examining OA had low median BLLs (<2.5 (.ig/dL). study participants
were born prior to the phase-out of leaded gasoline and therefore likely had much higher past Pb
exposures, making it difficult to characterize the specific timing, duration, frequency, and level of Pb
exposure that contributed to the observed associations. Additionally, similar to studies of osteoporosis,
the cross-sectional nature of the studies does not rule out the possibility that the association is driven by
cartilage turnover resulting in increased Pb in blood.
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9.5.3.3
Oral Health
Recent epidemiologic studies of Pb exposure and oral health are split into two major categories:
(1) periodontal disease in adults and (2) dental caries in children.
A limited number of recent studies of periodontal disease in adults examined overlapping
KNHANES cycles from 2008 to 2010 (Han et al.. 2013; Kim and Lee. 2013; Won et al.. 2013). These
studies, all of which defined periodontal disease according to the World Health Organization's
Community Periodontal Index, provided consistent evidence of an association between BLLs and the
prevalence of periodontitis. All of the studies included extensive adjustment for potential confounders,
including oral hygiene. Given that these studies examined largely overlapping study populations, the
observed results should not be considered independent evidence of an association. Kim and Lee (2013)
noted associations that were stronger in magnitude in men (OR =1.85 [95% CI: 1.26, 2.71] per two-fold
higher BLLs) compared to women (OR= 1.30 [95% CI: 0.88, 1.91] per two-fold higher BLLs), and that
associations were slightly attenuated, but still positive after adjustment for blood mercury (Hg) and
cadmium (Cd; 1.69 [95% CI: 1.15, 2.50] and 1.24 [95% CI: 0.83, 1.85], respectively). In stratified
analyses to examine effect modification by smoking, effect estimates were imprecise (i.e., wide 95% els),
but comparable in magnitude for smokers and non-smokers (Han et al.. 2013; Won et al.. 2013).
Recent epidemiologic studies of dental caries in children included more diverse study
populations. A prospective analysis of mother-child pairs that recruited from hospitals serving low- to
moderate-income populations in Mexico examined the relationship between Pb biomarkers at different
developmental windows and incidence of decayed, missing, and filled teeth (DMFT) in adolescence [10
to 18 years old; Wu et al. (2019)1. The authors reported 12 to 17% higher risk of DMFT associated with
higher prenatal and early childhood BLLs. No associations were observed with concurrent BLLs or
postnatal maternal bone Pb. Prenatal (mean: 5.24 to 6.36 (ig/dL) and early childhood (mean: 15.18 to
15.48 (ig/dL) BLLs were notably higher than concurrent levels (mean: 3.60-3.34 (.ig/dL). which is
consistent with age-specific patterns of Pb kinetics (Sections 2.2 and 2.4). Wu et al. (2019) additionally
stratified their models by sugar sweetened beverage intake (SSBI) and observed strong effect
modification, with stronger associations between prenatal and early childhood BLLs and DMFT score in
children with high SSBI. A prospective analysis in the same cohort reported positive but imprecise
associations between childhood Pb concentrations and DMFT in adolescence (Yepes et al.. 2020). In
recent cross-sectional studies with lower BLLs (see Table 9-6), BLLs in young children were associated
with increased prevalence of dental caries in deciduous teeth (Kim et al.. 2017; Wiener et al.. 2015). but
not permanent teeth (Kim et al.. 2017).
9.5.4 Toxicological Studies on the Musculoskeletal System
The 2013 Pb ISA (U.S. EPA. 2013) evaluated a number of toxicological studies that
demonstrated changes in bone cell function as a result of replacement of bone calcium with Pb depression
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in early bone growth. Studies also reported Pb-induced effects on cell proliferation, procollagen type I
production, intracellular protein, and osteocalcin in human dental pulp cell cultures. Earlier work,
summarized in the 2006 Pb AQCD (U.S. EPA. 2006). reported concentration-dependent depression of
early bone growth after gestational exposure of rodents to Pb. Recent evidence is limited. In a study of
lifetime Pb exposure in mice, Beier et al. (2016) reported a reduction in osteoclast activity and a
subsequent disruption in bone accrual in Pb-exposed mice. In another publication, the same group
reported no other musculoskeletal effects resulting from Pb exposure alone (Beier et al.. 2017).
9.5.5 Biological Plausibility
This section describes biological pathways that potentially underlie musculoskeletal effects of Pb.
Figure 9-2 depicts the proposed pathways as a continuum of upstream events, connected by arrows, which
may lead to downstream events observed in epidemiologic studies. This discussion of how exposure to Pb
may lead to musculoskeletal effects contributes to an understanding of the biological plausibility of
epidemiologic results evaluated above. Note that the structure of the biological plausibility sections and
the role of biological plausibility in contributing to the weight-of-evidence analysis used in the 2024 Pb
ISA are discussed in Section IS.4.2.
The proposed pathway, outlined in Figure 9-2, involves both direct and indirect effects of Pb that
could plausibly result in the weakening of bones and increased risk of fractures as well as the dental
effects that are measured in epidemiologic studies. Skeletal bone development and biomechanical
strength is controlled by the balance between osteoblasts, the cells responsible for the production of bone
matrix, and osteoclasts, the cells responsible for bone resorption. Dysregulation of this balance can lead to
bone loss and decreased mineralization. Pb can directly replace Ca2+ in the bone matrix as well as exert
direct effects on bone cells to alter bone development. Pb can also alter bone growth and differentiation
signals that can further disrupt the balance of bone formation and resorption.
As discussed in the 2013 Pb ISA, Pb suppresses the differentiation of osteoblasts and promotes
osteoclast function which could result in delayed bone development and reduced bone mechanical
integrity. Recent literature supports this hypothesis as studies have continued to show that animals treated
with Pb have decreased bone mineralization (Li et al.. 2020a; Sheng et al.. 2020; Oi et al.. 2019;
Olchowik et al.. 2014). bone weight (Alvarez-Lloret et al.. 2017; de Figueiredo et al.. 2014). and reduced
trabecular bone (Li et al.. 2020a; Sheng et al.. 2020; Alvarez-Lloret et al.. 2017; Beier et al.. 2017). Many
of these studies show concurrent changes in osteoblastic and osteoclastic markers that support an overall
shift to increased bone resorption. For example, recent in vivo studies have seen reductions markers of
osteoblast differentiation (Qi et al.. 2019; Zhang et al.. 2019; Beier et al.. 2017). reductions of proteins
that suppress osteoclast activity (Li et al.. 2020a; Sheng et al.. 2020; Oi et al.. 2019; Kupraszewicz and
Brzoska. 2013). and increases of markers of osteoclast activity (Li et al.. 2020a; Qi et al.. 2019; Zhang et
al.. 2019; Kupraszewicz and Brzoska. 2013) suggesting that bone changes result from dysregulation of
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the balance between bone formation and bone resorptive processes. The mechanism behind the reduced
osteoblastic activity is not fully understood but both direct and indirect mechanisms have been proposed.
Support for a direct action of Pb on osteoblast function comes from in vitro studies showing that
Pb treatment of primary osteoblasts leads to reduction in mineral deposition (Beier et al., 2015; Abbas et
al., 2013; Ma et al., 2012). Previously reviewed data also implicated changes in TGF|3, bone morphogenic
protein (BMP), nuclear factor kappa B (NF-kB), and activator protein-1 signaling (U.S. EPA, 2013).
Recent studies suggest that Pb-induced suppression of Wnt signaling and upregulation of the protein
sclerostin may also be involved (Sun et al.. 2019; Beier et al., 2017; Beier et al., 2015). Similar studies of
dental pulp cultures showed that in vitro treatment with Pb resulted in decreased cell proliferation and
reduced extracellular matrix deposition. This could explain the increased incidence of dental carries in
epidemiology studies.
Indirect mechanisms of Pb treatment have also been discussed in the 2013 Pb ISA. The
replacement of Pb for Ca2+ in cells can lead to Ca2+ release. The 2013 Pb ISA and 2006 Pb AQCD
discussed studies that found that Pb treatment leads to increased systemic Ca2+ levels in the blood stream
(U.S. EPA, 2013, 2006). Calcium is a cellular signaling molecule involved in mitochondrial function and
cell death and thus changes in calcium signaling could have effects on cells elsewhere in the body. Bone
growth can be affected by systemic signaling of hormones and vitamins that regulate osteoblast formation
as well as storage and release of Ca2+ including parathyroid hormone (PTH), GH, BMP, and vitamin D.
As discussed previously in the 2006 Pb AQCD and the 2013 Pb ISA, Pb exposure can alter these pro-
osteoblastic signals which are thought to be involved in the reduction of bone growth and mineralization
seen following Pb exposure. Recent studies show similar alterations in calcitropic and osteoplastic signals
that could be responsible for reduced bone formation (Zhang et al., 2019; Kupraszewicz and Brzoska,
2013). Together, these data provide plausible indirect pathway by which Pb exposure can regulate skeletal
bone homeostasis.
The pathway for development of osteoarthritis is less well studied. Osteoarthritis results from
erosion of cartilage and articular bone in the joints. Chondrocytes are responsible for matrix deposition
and joint maintenance. Signaling though TGF|3 is thought to be important in proper joint maintenance. A
recent study showed that Pb treatment in rats induced cartilage loss which was associated with loss of
extracellular matrix proteins (Holz et al., 2012). In the same study, in vitro treatment of chondrocytes
from rat or chicks resulted in reduced markers of TGF|3 signaling and increased markers of matrix
degradation. These data suggest that Pb-induced osteoarthritis could be a result of Pb effects of
chondrocytes and subsequent cartilage degradation.
Teeth do not undergo the same bone turnover processes as skeletal bone and thus Pb incorporated
into the teeth is permanently sequestered. As discussed in the 2013 Pb ISA, dental effects of Pb are
thought to arise from the effects of Pb on enamel producing cells in combination with the incorporation of
Pb into areas of mineralization (U.S. EPA, 2013). Previously evaluated studies showed decrease cell
proliferation, procollagen type I production, intracellular protein, and osteocalcin in human dental pulp
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cell cultures (U.S. EPA. 2013). A recent study supports the link between Pb exposure and dental effects
by showing reduced molar diameter and increased dental cracks in the offspring of rats treated with Pb
during either gestation or lactation (Chen et al.. 2012). Together Pb-induced dental effects could result
from effects on dental pulp cells resulting in reduced matrix proteins.
The toxicologic data support Pb-induced alterations in multiple aspects of bone, teeth, and joint
maintenance. For skeletal bones, shift in the balance between bone building osteoblasts and bone
resorbing osteoclasts could be responsible for delayed bone growth and increased bone degeneration seen
in epidemiologic studies. In teeth and joints, Pb appears to suppress the synthesis of cellular matrix
proteins important for joint maintenance and enamel formation which could plausibly contribute to the
osteoarthritic and dental effects seen in some epidemiology studies.
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Pb
Exposure
Depressed cell growth
and mineralization
Altered
osteoblast/osteoclast
balance
Osteoporosis
Increased
falls/fractures
Depressed protein
synthesis
Osteoarthritis
Dental effects/Tooth
loss
Note: The boxes above represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and the arrows indicate a proposed relationship
between those effects. Solid arrows denote evidence of essentiality as provided, for example, by an inhibitor of the pathway used in an experimental study involving Pb exposure.
Dotted arrows denote a possible relationship between effects. Shading around multiple boxes is used to denote a grouping of these effects. Arrows may connect individual boxes,
groupings of boxes, and individual boxes within groupings of boxes. Progression of effects is generally depicted from left to right and color coded (white, exposure; green, initial effect;
blue, intermediate effect; orange, effect at the population level or a key clinical effect). Here, population-level effects generally reflect results of epidemiologic studies. When there are
gaps in the evidence, there are complementary gaps in the figure and the accompanying text below. IS.7.2 discusses the structure of the biological plausibility sections and the role of
biological plausibility in contributing to the weight-of-evidence analysis used in the 2022 Pb ISA.
Figure 9-2
Potential biological pathways for musculoskeletal effects following exposure to Pb.
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9.5.6
Summary and Causality Determination
The 2013 Pb ISA concluded that evidence was sufficient to determine "that a causal relationship
is likely to exist between Pb exposure and effects on bone and teeth" (U.S. EPA, 2013). This causality
determination was based on a small body of epidemiologic evidence showing associations between Pb
biomarker levels and effects on bones after adjusting for potential confounding by age and SES-related
factors, as well as strong toxicological evidence that reported effects on bone in animals following Pb
exposure. Specifically, a few epidemiologic studies indicated an association between higher Pb biomarker
levels and lower bone density in adults. A prospective study of older women provided evidence that
higher BLLs (>4 (ig/dL versus <3 (ig/dL) were associated with greater risk of falls and osteoporosis-
related fractures, as well as lower bone density measured after 2-4 years (Khalil et al., 2009). This finding
was supported by cross-sectional associations between higher BLLs and lower BMD (Campbell and
Auinger, 2007) and biochemical biomarkers of higher bone turnover (Nelson et al., 201 la; Machida et al.,
2009) in adults. In evaluating the cross-sectional epidemiologic evidence, it is difficult to determine
whether an increase in BLLs results from lower bone density or from higher bone turnover, and whether
either of these effects lead to a greater release of Pb from bone into the bloodstream. Exposure of animals
to Pb during gestation and the immediate postnatal period was reported to significantly depress early bone
growth with the effects showing concentration-dependent trends. Systemic effects of Pb exposure
included disruption in bone mineralization during growth, alteration in bone cell differentiation and
function due to alterations in plasma levels of growth hormones and calcitropic hormones such as 1,25-
[OH]2D3 and impact on Ca2+- binding proteins and increases in Ca2+ and phosphorus concentrations in
the bloodstream. Bone cell cultures exposed to Pb had altered vitamin D-stimulated production of
osteocalcin accompanied by inhibited secretion of bone-related proteins such as osteonectin and collagen.
In addition, Pb exposure caused suppression in bone cell proliferation most likely due to interference from
factors such as GH, EGF, transforming growth factor-beta 1 (TGF-|31), and PTHrP.
In addition to effects on bone, epidemiologic and toxicological studies evaluated in the 2013 Pb
ISA provided evidence of Pb-related effects on teeth. A limited number of epidemiologic studies reported
associations between higher BLLs and a higher prevalence of dental caries in children (Moss et al., 1999)
and periodontitis in adults (Saraiva et al., 2007). Additionally, higher patella and tibia Pb levels were
associated with tooth loss in men participating in the NAS (Arora et al., 2009). This epidemiologic
evidence was based on cross-sectional analyses, which precludes conclusions about the directionality of
effects. However, these findings are supported by toxicological evidence in animals for Pb-induced
increases in Pb uptake into teeth and decreases in cell proliferation, procollagen type I production,
intracellular protein, and osteocalcin in cells exposed to Pb in vitro. Despite evidence for associations
between Pb exposure and effects in bone and teeth at relatively low concurrent BLLs, these outcomes
were most often examined in older adults that have been exposed to higher levels of Pb earlier in life.
Therefore, uncertainty still remains concerning the Pb exposure level, timing, frequency, and duration that
contribute to the observed associations.
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Recent cross-sectional epidemiologic studies continue to support associations between Pb
exposure and effects on bone. The majority of recent studies of osteoporosis or osteopenia were
conducted in female populations or included models stratified by sex to account for sex-specific
difference in osteoporosis and osteopenia incidence. The evaluated studies provide generally consistent
evidence of a positive association between low BLLs (mean/median ranges cross studies: 1.03 to
3.4 (ig/dL) and osteoporosis or osteopenia in women (Li et al.. 2020b; Wang et al.. 2019; Pollack et al..
2013; Cho et al.. 2012; Lee and Kim. 2012). Other studies also observed positive associations in models
including men and women (Lee and Park. 2018; Lim et al.. 2016). but the inferences that can be drawn
from these studies are limited due to the previously noted sex differences in BMD. A few recent cross-
sectional studies also reported associations between low BLLs and symptomatic and radiographic OA in
the knee (Park and Choi. 2019; Nelson et al.. 2011b). These findings were supported by another study
demonstrating associations between BLLs and some biomarkers of joint tissue metabolism, which could
either lead to OA or be indicative of prevalent OA (Nelson et al.. 201 la). These studies of OA represent
an emerging area of research for an endpoint that was not discussed in the 2013 Pb ISA. Recent
epidemiologic evidence is prone to similar uncertainties and limitations identified in the 2013 Pb ISA.
Notably, the cross-sectional design of these studies does not establish temporality between the exposure
and outcome. This may be particularly relevant for health outcomes that correlate with bone turnover
rates that could lead to higher BLLs. Additionally, although a number of recent studies have been
conducted in adult populations with low BLLs, uncertainty regarding past exposures continues to limit the
characterization of the Pb exposure levels, timing, frequency, and duration that contribute to the observed
associations.
The recent toxicological evidence base for effects on bones is smaller, but consistent with
findings from the 2013 Pb ISA and coherent with recent epidemiologic evidence. Notably, a recent study
reported a reduction in osteoclast activity and a disruption in bone accrual in Pb-exposed animals (Beier
et al.. 2016). This finding, along with similar evidence from previous ISAs and AQCDs, provides support
for a temporal relationship between Pb exposure and effects on bone accrual and bone density that cannot
be established by the available cross-sectional epidemiologic evidence.
In addition to studies of Pb exposure and effects on bone, recent epidemiologic studies have also
explored the relationship between BLLs and effects on teeth. Recent studies in adults focused on the
prevalence of periodontitis, whereas studies in children examined the prevalence or incidence of dental
caries. A group of studies examining overlapping KNHANES cycles observed positive associations
between low BLLs and periodontitis prevalence in adults (Han et al.. 2013; Kim and Lee. 2013; Won et
al.. 2013). including some evidence of a stronger association in men, and persistent associations in models
adjusting for Hg and Cd (Kim and Lee. 2013). Given the largely overlapping study populations, the
observed results should not be interpreted as independent evidence of an association. Additionally, the use
of BLLs in adult populations with higher past exposures limits the ability to characterize the Pb exposure
levels, timing, frequency, and duration that contribute to the observed associations. In a prospective birth
cohort study of low- to moderate-income mother-child pairs, increases in prenatal and early childhood
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BLLs were associated with increased risk of dental caries in adolescence (Wu et al.. 2019). The authors
also observed a null association with concurrent BLLs, which suggests that there may be critical windows
of exposure earlier in life. These findings were supported by a few cross-sectional studies that reported
associations between BLLs in early childhood and increased prevalence of dental caries in deciduous
teeth (Kim et al.. 2017; Wiener et al.. 2015). No recent toxicological studies have examined the effects of
Pb exposure on teeth, but as described earlier, previous and recent mechanistic evidence provides
biological plausibility for the observed epidemiologic associations.
Overall, the collective evidence is sufficient to conclude that there is likely to be a causal
relationship between Pb exposure and musculoskeletal effects. This causality determination is based
on an expanded epidemiologic evidence base that continues to demonstrate associations between BLLs
and various musculoskeletal effects after adjusting for potential confounding, as well as strong
toxicological evidence for effects on bone in animals following Pb exposure. Although the recent
epidemiologic evidence is consistent with the findings highlighted in the 2013 Pb ISA, recent studies do
not thoroughly address uncertainties identified in the 2013 Pb ISA, including unclear temporality of
exposure and outcome resulting from mostly cross-sectional study designs. This uncertainty is
particularly important for studies examining BMD and osteoporosis due to the possibility that
associations could be driven by increased BLLs due to higher bone turnover in individuals with low BMD
or osteoporosis. Although there are not many recent toxicological studies that meet PECOS relevance, the
evaluated studies are consistent with a large evidence base from the 2013 Pb ISA and AQCD, which
provides support for the observed epidemiologic associations. The key evidence, as it relates to the causal
framework, is summarized in Table 9-2.
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Table 9-2 Summary of evidence for a likely to be causal relationship between Pb exposure and
musculoskeletal effects
Rationale for Causality
Determination3
Key Evidence13
Key References'3
Pb Biomarker Levels Associated with
Effects0
Consistent evidence from
epidemiologic studies of
osteoporosis and osteopenia
Evidence from cross-sectional
epidemiologic studies supports
associations between Pb exposure and
osteoporosis or osteopenia in adult
female populations.
Cho et al. (2012)
Wanq et al. (2019)
Lee and Kim (2012)
Pollack et al. (2013)
Li et al. (2020b)
Mean/median BLL ranges cross studies:
1.03 to 3.4 |jg/dL
Supporting evidence from
toxicological studies with
relevant exposures
investigating effects on bone
Toxicological evidence in rodents is
coherent with epidemiologic evidence
and provides support for a temporal
relationship between Pb exposure and
effects on bone accrual and bone density
Beier et al. (2016)
(U.S. EPA. 2013)
(U.S. EPA. 2006)
Mean BLL range of 20.8 to 49.9 |jg/dL
Consistent evidence from
epidemiologic studies of
dental caries in children
A prospective birth cohort study provides
evidence that increases in prenatal and
early childhood BLLs are associated with
increased risk of dental caries in
adolescence
Wu et al. (2019)
Mean BLLs (males, female):
15.48, 15.18 |jg/dL
Supporting cross-sectional evidence of
associations between early childhood
BLLs and dental caries in deciduous
teeth
Kim et al. (2017)
Wiener et al. (2015)
Geometric Mean BLLs: 1.53 |jg/dL
Mean NR (28.2% <2 pg/dL; 48.3% 2 to
<5 ug/dL; 18.4% 5 to <10 ug/dL; 5.1%
>10 pg/dL)
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Rationale for Causality
Determination3
Key Evidence13
Key References'3
Pb Biomarker Levels Associated with
Effects0
Biological Plausibility Pb can directly replace Ca2+ in the bone Section 9.5.4.
matrix as well as exert direct effects on
bone cells to alter bone development. Pb
can also alter bone growth and
differentiation signals that can further
disrupt the balance of bone formation
and resorption. Pb has also been shown
to decrease cell proliferation, procollagen
type I production, intracellular protein,
and osteocalcin in human dental pulp cell
cultures.
Preamble to the ISAs (U.S. EPA. 2015).
where applicable, to uncertainties or
BLLs = blood lead levels; Ca2+ = calcium ion(s); NR = not reported; Pb = lead.
aBased on aspects considered in judgments of causality and weight-of-evidence in causal framework in Table I and Table II of the
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and,
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
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9.6
Effects on Ocular Health
9.6.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the Current
Review
This section of effects on ocular health focuses on impairments related to the structure of the eye,
including but not limited to cataracts, glaucoma, macular degeneration, and retinal stippling. Studies
examining effects on vision that are related to sensory processing in the central nervous system can be
found in Appendix 3 of the 2024 Pb ISA (Sections 3.5.6.2 and 3.6.3.2). The 2013 Pb ISA concluded that
because the studies of effects on ocular health were of insufficient quantity and quality, the overall
evidence was "inadequate to determine a causal relationship between Pb exposure and ocular effects"
(U.S. EPA. 2013). There were very few studies evaluated in the 2013 Pb ISA that examined Pb exposure
and ocular effects in humans or animals. Those studies that were reviewed examined disparate outcomes
and the epidemiologic studies lacked rigorous statistical analyses.
9.6.2 Scope
The scope of this section is defined by PECOS statements. The PECOS statement defines the
objectives of the review and establishes study inclusion criteria thereby facilitating identification of the
most relevant literature to inform the Pb ISA.27 In order to identify the most relevant literature, the body
of evidence from the 2013 Pb ISA was considered in the development of the PECOS statements for this
Appendix. Specifically, well-established areas of research; gaps in the literature; and inherent
uncertainties in specific populations, exposure metrics, comparison groups, and study designs identified
in the 2013 Pb ISA inform the scope of this Appendix. The 2013 Pb ISA used different inclusion criteria
than the 2024 Pb ISA, and the studies referenced therein often do not meet the current PECOS criteria
(e.g., due to higher or unreported biomarker levels). Studies included in the 2013 Pb ISA, including many
that do not meet the current PECOS criteria, are discussed in this appendix to establish the state of the
evidence prior to this assessment. Except for supporting evidence used to demonstrate the biological
plausibility of Pb-associated effects on the ocular health, recent studies were only included if they
satisfied all of the components of the following discipline-specific PECOS statements:
27The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
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Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect.
Exposure: Exposure to Pb28 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure;29 or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles).
Outcome: Effects on ocular health.
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages).
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a BLL of 30 (ig/dL or below.31 ¦31
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control.
Outcomes: Ocular effects.
Study design: Controlled exposure studies of animals in vivo.
28Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area of particular
relevance to the NAAQS review (e.g., longitudinal studies designed to examine recent versus historical Pb
exposure).
29 Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 (im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 |im3 (PM2.5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is difficult
to assess the representativeness of these concentrations to population exposure [Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PNLs with BLLs are lacking.
3l Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
31This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLL.
The 95th percentile of the 2011-2016 NHANES distribution of BLL in children (1-5 years; n= 2,321) is 2.66 (ig/dL
(Egan et al„ 2021) and the proportion of individuals with BLL that exceed this concentration varies depending on
factors including (but not limited to) housing age, geographic region, and a child's age, sex, and nutritional status.
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9.6.3
Epidemiologic Studies on Ocular Health
A limited number of epidemiologic studies evaluated in the 2006 Pb AQCD (U.S. EPA, 2006)
and 2013 Pb ISA (U.S. EPA, 2013) provide some evidence of an association between exposure to Pb and
ocular health. As part of the longitudinal Normative Aging Study, Schaumbcrg et al. (2004) analyzed
prevalence of cataracts in adult males in relation to blood Pb, patella bone Pb, or tibia bone Pb levels.
Covariate-adjusted odds ratios for cataracts were elevated for the highest quintiles of tibia (3.19 [95% CI:
1.48, 6.90]) and patella (1.88 [95% CI: 0.88, 4.02]) Pb levels compared to the lowest. A null association
was observed for the highest quintile of blood Pb (0.89 [95% CI: 0.46, 1.72]). This may suggest a role for
past and cumulative long-term exposures, which aligns with the chronic nature of cataracts. Evidence for
other ocular diseases from less robust studies provided inconclusive evidence due to study limitations. A
cross-sectional study of macular degeneration reported higher concentrations of Pb in the retinal tissue of
donors with macular degeneration compared to those without (Erie et al., 2009). However, the authors did
not control for confounders in this comparison of means. Another study measured BLLs in smokers and
non-smokers with cataracts, but the authors did not make comparisons between exposure to Pb and
severity of cataracts (Mosad et al., 2010).
Recent studies provide inconsistent evidence of an association between exposure to Pb and ocular
effects. The majority of recent studies evaluating ocular health and Pb exposures are population-based
cross-sectional analyses, which are unable to establish temporality between exposure and outcome.
Additionally, because many of the observed ocular impairments generally occur in older adult populations
who likely had higher past than current Pb exposure, there is uncertainty regarding the Pb exposure level,
duration, frequency, and timing that may contribute to any observed associations. Measures of central
tendency for blood and/or bone Pb levels used in each study, along with other study-specific details,
including study population characteristics and select effect estimates, are highlighted in Table 9-13. An
overview of the recent evidence is provided below.
A limited number of recent studies have evaluated the relationship between levels of Pb in the
blood or bone and glaucoma. The strongest evidence for an association comes from a longitudinal
analysis of the Veterans Affairs NAS, a prospective cohort study of male Veterans (Wang et al„ 2018b).
Wang et al. (2018b) reported that higher tibia and patella Pb were associated with 28% (95% CI: -1%,
65%) and 42% (95% CI: 11%, 82%) higher risk of primary open-angle glaucoma, respectively. These
results are supported by a recent KNHANES mediation analysis that evaluated intraocular pressure,
which is an important risk factor for glaucoma (Park and Choi, 2016). The authors reported that each
1 (ig/dL higher BLL was associated with 0.09 mmHg (95% CI: 0.06, 0.12 mmHg) higher intraocular
pressure, after accounting for indirect effects of exposure to Pb through higher blood pressure. The
estimated total effect (i.e., not controlling for mediation by blood pressure) for a 1 (ig/dL higher level of
blood Pb was 0.11 mmHg (standard error not reported). In contrast, two recent large cross-sectional
studies of the KNHANES did not observe an association between BLLs and glaucoma (Lee et al., 2016;
Lin et al„ 2015). However, potential associations with chronic age-related diseases, such as glaucoma,
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may be better evaluated using measurements of Pb in bone, which has a much longer half-life than in
blood and is therefore a better indicator of cumulative exposure.
In addition to studies of glaucoma, there were also a few recent population-based cross-sectional
studies that examined the association between BLLs and age-related macular degeneration (AMD) in
older adults (Hwang et al.. 2015; Park et al.. 2015; Wu et al.. 2014). AMD is a common eye-disorder in
older adults that is caused by retinal damage, resulting in deteriorated central vision. Two recent studies
of the KNHANES provided evidence of an association between BLLs and AMD (Hwang et al.. 2015;
Park et al.. 2015). Using data from the 2008-2011 cycles of KNHANES, Park et al. (2015) reported 12%
(95% CI: 2%, 23%) higher odds of early-stage AMD (i.e., damaged macula with no vision loss) and 25%
(95% CI: 5%, 50%) higher odds of late-stage AMD (i.e., damaged macula with vision loss) per 1 (ig/dL
higher BLLs. In a similar study that analyzed one additional year of KNHANES data (2008-2012),
Hwang et al. (2015) similarly observed higher odds of early-stage AMD corresponding to higher quintiles
of Pb exposure. Notably, in stratified analyses examining effect modification by sex, the observed
associations in the total population appeared to be driven by a much stronger association in women. The
authors also reported associations for late-stage AMD, but the case numbers were so low for each quintile
that the reduced statistical power to detect an association made the results unreliable. In contrast to the
results from the KNHANES studies, Wu et al. (2014) reported null associations between BLLs and AMD
in an analysis of older adults in the 2005-2008 cycles of the U.S. NHANES.
Additional cross-sectional studies examined other ocular health effects for disparate outcomes,
including an NHANES analysis of cataract surgery in older adults (Wang et al.. 2016) and a KNHANES
study of dry eye disease (Jung and Lee. 2019). Both of these studies reported null associations between
BLLs and the ocular health outcome of interest.
9.6.4 Toxicological Studies on Ocular Health
The 2013 Pb ISA (U.S. EPA. 2013) made note of a limited number of animal studies finding Pb-
induced mouse retinal progenitor cell proliferation and neurogenesis, as well as increased opacity of rat
lens after Pb exposure.
Two recent toxicological studies were identified since the 2013 Pb ISA for inclusion in the 2024
Pb ISA. Perkins et al. (2012) described remodeling of rod and cone synaptic mitochondria in mice after
postnatal exposure to Pb acetate in drinking water (21 (ig/dL BLL at weaning). The observed Pb-induced
changes are consistent with deficits in range of vision. The effect of Pb on rod and cone mitochondria was
mediation by Bcl-xL, a protein that has been implicated in Pb-induced apoptosis. Using adult rats exposed
to Pb acetate in drinking water (1-20 (ig/dL BLL), Shen et al. (2016) found increased blood-retinal
permeability. The authors noted an association between long-term increased vascular permeability with
retinal dysfunction and degeneration.
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9.6.5
Summary and Causality Determination
The 2013 Pb ISA concluded that evidence was "inadequate" to determine a causal relationship
between Pb exposure and ocular health effects (U.S. EPA, 2013). This causality determination was based
on an insufficient quantity and quality of studies in the cumulative body of evidence. Although a cross-
sectional epidemiologic study reported higher concentrations of Pb in the retinal tissue of donors with
macular degeneration compared to those without (Erie et al., 2009), the study did not account for smoking
status as potential confounder. Toxicological studies were limited in number, but reported Pb-induced
retinal progenitor cell proliferation, retinal electroretinograms, and lens opacity.
Since the release of the 2013 Pb ISA, there has been an increase in the number of epidemiologic
studies that examine the relationship between Pb exposure and ocular health effects. Recent
epidemiologic studies provide inconsistent evidence of an association between Pb exposure and ocular
health effects. The strongest evidence comes from a prospective cohort study of male Veterans that
reported large, but imprecise associations between bone Pb levels and glaucoma (Wang et al., 2018b).
These results are supported by a cross-sectional association between BLLs and intraocular pressure,
which is an important risk factor for glaucoma (Park and Choi, 2016). However, additional population-
based cross-sectional studies in the same population reported null associations between BLLs and
glaucoma (Lee et al„ 2016; Lin et al„ 2015). No recent experimental studies examined endpoints related
to glaucoma.
Findings from a limited number of population-based cross-sectional studies of Pb exposure and
AMD were inconsistent across populations - with null results observed in a U.S.-based study and a
positive association in a South Korean-based study. A recent toxicological study reported Pb-induced
increases in blood-retinal permeability, which may lead to increased risk of macular degeneration.
Although the evidence base has expanded since the completion of the 2013 Pb ISA, the limited
number of studies and the inconsistent results do not provide sufficient information to draw a conclusion
regarding causality. Thus, the evidence remains inadequate to infer the presence or absence of a
causal relationship between Pb exposure and ocular health effects.
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9.7
Effects on the Respiratory System
9.7.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the
Current Review
The 2013 Pb ISA evaluated studies of respiratory effects related to inflammatory and atopic
diseases (like asthma) separately from effects on lung function, morphology, and respiratory symptoms.
Similarly, in this review, studies evaluating the effect of Pb on asthma are discussed with effects on the
immune system in Appendix 6. This section discusses the effects of Pb on the respiratory system in the
otherwise healthy lung. The 2013 Pb ISA concluded that there was "insufficient quantity and quality of
studies" related to the impacts of Pb on the non-asthmatic lung and the evidence was therefore
"inadequate to determine a causal relationship" (U.S. EPA. 2013). Epidemiologic studies in non-
asthmatics were lacking in number, consistency, and statistical rigor, despite observed associations
between BLLs and respiratory effects in children and asthmatics (Appendix 6). The few respiratory
toxicological studies described previously were in vivo and in vitro studies that administered concentrated
ambient particulate matter, of which Pb was a component. The ability to evaluate the independent effect
of Pb in these studies was limited due to the inability to account for confounding effects of co-pollutants
and the lack of characterization of Pb particles in the samples. Given the limitations of these studies, the
scope for this review was narrowed to remove toxicological studies that analyzed the health effects of Pb
containing mixtures but lacked a Pb alone treatment group.
9.7.2 Scope
The scope of this section is defined by PECOS statements. The PECOS statement defines the
objectives of the review and establishes study inclusion criteria thereby facilitating identification of the
most relevant literature to inform the Pb ISA.32 In order to identify the most relevant literature, the body
of evidence from the 2013 Pb ISA was considered in the development of the PECOS statements for this
Appendix. Specifically, well-established areas of research; gaps in the literature; and inherent
uncertainties in specific populations, exposure metrics, comparison groups, and study designs identified
in the 2013 Pb ISA inform the scope of this Appendix. The 2013 Pb ISA used different inclusion criteria
than the 2024 Pb ISA, and the studies referenced therein often do not meet the current PECOS criteria
32The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
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(e.g., due to higher or unreported biomarker levels). Studies included in the 2013 Pb ISA, including many
that do not meet the current PECOS criteria, are discussed in this appendix to establish the state of the
evidence prior to this assessment. Except for supporting evidence used to demonstrate the biological
plausibility of Pb-associated effects on the immune system, recent studies were only included if they
satisfied all of the components of the following discipline-specific PECOS statements:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect.
Exposure: Exposure to Pb33 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure34; or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles).
Outcome: Effects on the respiratory system.
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages).
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a BLL of 30 (ig/dL or below.35,36
33Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area of particular
relevance to the NAAQS review (e.g., longitudinal studies designed to examine recent versus historical Pb
exposure).
34Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 |im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 |im3 (PM2.5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is difficult
to assess the representativeness of these concentrations to population exposure [Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PNLs with BLLs are lacking.
35Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
36This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLL.
The 95th percentile of the 2011-2016 NHANES distribution of BLL in children (1-5 years; n= 2,321) is 2.66 (ig/dL
(Egan et al„ 2021) and the proportion of individuals with BLL that exceed this concentration varies depending on
factors including (but not limited to) housing age, geographic region, and a child's age, sex, and nutritional status.
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Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control.
Outcomes: Effects on the respiratory system.
Study design: Controlled exposure studies of animals in vivo.
9.7.3 Epidemiologic Studies on the Respiratory System
A limited number of epidemiologic studies evaluated in the 2013 Pb ISA did not provide strong
evidence of an association between BLLs and airway responses in asthma-free populations. Further, these
studies lacked rigorous statistical analysis and included limited consideration of potential confounders. In
panel and time-series epidemiologic studies considering ambient air Pb (measured in PM2.5 or PM10 air
samples), associations were reported between short-term increases in air Pb and decreases in lung
function and increases in respiratory symptoms and asthma hospitalizations in children but not adults.
Despite this evidence for respiratory effects related to air Pb concentrations, the limitations of air Pb
studies - including the limited data on the size distribution of Pb-PM, the uncertain relationships of Pb-
PM10 and Pb-PMo.s with BLLs, and the lack of adjustment for other correlated particulate matter (PM)
chemical components - precluded firm conclusions about ambient air Pb-associated respiratory effects.
Recent studies have examined lung function and respiratory symptoms in non-asthmatic children and
adults. While the majority of recent studies utilized cross-sectional designs that are unable to establish
temporality between exposure and outcome, most adjust for a wide range of potential confounders and
examine populations with lower BLLs. In general, recent evidence in children is inconsistent, though
there is some evidence from a prospective cohort study that BLLs are associated with accelerated lung
function decline in adults. Notably, because adult populations likely had higher past than current Pb
exposure, there is uncertainty regarding the Pb exposure level, duration, frequency, and timing that may
contribute to the observed association. Measures of central tendency for blood and/or serum Pb levels
used in each study, along with other study-specific details, including study population characteristics and
select effect estimates, are highlighted in Table 9-15. An overview of the recent evidence, delineated by
lifestage, is provided below.
9.7.3.1 Respiratory Effects in Children
A limited number of recent cross-sectional studies have examined the relationship between BLLs
and pulmonary function or respiratory symptoms in children. Studies conducted in different locations
reported inconsistent evidence of an association between BLLs and pulmonary function. In an analysis of
6- to 17-year-old children participating in the 2011-2012 NHANES survey cycle, Madrigal et al. (2018)
reported modest and imprecise positive associations between BLLs and mean forced expiratory volume
(FEV1) (41.9 mL [95% CI: -46.9, 130.6 mL]) and forced vital capacity (FVC) (45.5 mL [95% CI: -49.2,
140.2 mL]) when comparing children with BLLs in the highest quartile (>0.86 (ig/dL) to children with
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BLLs in the first quartile (<0.44 (ig/dL). Similar comparisons were null for FEV1:FVC and forced
expiratory flow (FEF)25%-75%. Notably, the study population had a very low median BLL (0.56 (.ig/dL).
and there were small exposure contrasts between exposure quartiles, which may have limited the
statistical power to detect an association. In contrast with the NHANES analysis, smaller cross-sectional
studies conducted in preschool-aged children in China (Zeng et al.. 2017) and 10- to 15-year-old children
in Poland (Little et al.. 2017) observed limited evidence of associations between BLLs and decreased
FVC (Little et al.. 2017; Zeng et al.. 2017) or FEV1 (Zeng et al.. 2017). Both studies noted small and
imprecise associations and had small sample sizes. Limited statistical power resulting from a small
sample size reduces the likelihood of detecting a true effect and lowers precision, which might explain the
incongruous results. Additionally, the associations observed by Little et al. (2017) may have been subject
to unmeasured confounding (e.g., by age, SES factors, environmental tobacco smoke), as the authors only
adjusted their regression models for children's heights.
In addition to studies of pulmonary function, a single study examined respiratory symptoms in
children. Zeng et al. (2016) reported inconsistent associations between BLLs and respiratory symptoms in
preschool-aged children in China, including some living in a community near an e-waste facility. The
authors compared children with BLLs >5 (ig/dL to those with BLLs <5 (ig/dL and reported that those in
the higher exposure group had lower odds of parental-reported wheeze and dyspnea, slightly higher odds
of parental-reported phlegm, and no perceptible difference in parental-reported cough. Caution is
warranted in interpreting results of parental-reported symptoms in locations with known environmental
contamination due to potential over-reporting of symptoms.
9.7.3.2 Respiratory Effects in Adults
A limited number of recent studies have examined the relationship between blood or serum Pb
levels and respiratory effects in adults. There is evidence from a prospective cohort study that BLLs are
associated with accelerated lung function decline in adults, although a large, population-based cross-
sectional study reports conflicting results. All of the studies evaluated in this subsection reported low
levels of blood or serum Pb levels (mean and geometric mean levels <3 (.ig/dL).
The most compelling evidence of an association between Pb exposure and lung function in adults
comes from a prospective cohort study of adults living adjacent to a large industrial complex in South
Korea (Pak et al.. 2012). The authors reported that BLLs were associated with accelerated lung function
decline, measured as the difference in spirometry measurements taken at baseline and after two-years of
follow-up. Specifically, Pak et al. (2012) noted accelerated decline in FVC (-177 mL [95% CI: -330,
-24]) and FEV1 (-107 mL [95% CI: -215, 1]) per 1 (ig/dL higher BLL at baseline. Notably, because
adult populations likely had higher past than current Pb exposure, there is uncertainty regarding the Pb
exposure level, duration, frequency, and timing that may contribute to the observed association. In
contrast to results from Pak et al. (2012). a recent cross-sectional study of 2008-2012 KNHANES
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participants with low BLLs observed null associations between BLLs and FVC and FEV1 in adults
(Leem et al.. 2015).
Leem et al. (2015) also examined obstructive lung function (FEV1/FVC <0.7) in the same
population and observed a null association with BLLs. In a similar recent analysis of a large population-
based health survey (NHANES), (Rokadia and Agarwal. 2013) reported a large, but imprecise association
between serum Pb levels and obstructive lung function (OR= 1.94 [95%: 1.10, 3.42] per 1 (ig/dL higher
level of serum Pb) that appears to be driven by an association in participants with moderate to severe
obstructive lung function (OR = 3.49 [95%: 1.70,7.15] perl (ig/dL higher level of serum Pb). The
observed associations were similar in analyses stratified by smoking status, although the associations in
non-smokers were even less precise due to a smaller number of cases.
9.7.4 Toxicological Studies on the Respiratory System
The 2013 Pb ISA evaluated a limited number of studies investigating the effects of ambient
particulate mixtures of which Pb was a component. The effects directly attributable to Pb were not able to
be distinguished from other confounding mixture components. The PECOS criteria used in the 2024 Pb
ISA to identify new respiratory toxicological studies focused on identifying studies that studied Pb
exposure alone. One study reviewed in the 2013 Pb ISA showed that injection of Pb acetate resulted in
histologic signs of damage and inflammation in the lung although uncertainty regarding the biological
relevance of Pb injection remained A few new experimental studies were identified that investigated the
effect of inhaled Pb and met our PECOS criteria (Table 9-9). The studies, all published by the same
group, assessed the localization and clearance of inhaled ultrafine (>100 nm in diameter) Pb particles and
the corresponding effect on lung (and secondary organ) tissue structure. These studies involved 2-
11 weeks of exposures (24 hours/day, 7 days/week) to inhaled Pb nanoparticles after which the
investigators analyzed lung histology and markers of lung damage. Exposure of female mice to roughly
106particles/cm3 lead oxide (PbO) particles for 6 weeks led to a mean BLL of 132 ng/g (-13.922 (ig/dL)
and corresponded to histological signs of lung damage including alveolar septal wall thickening,
emphysema, perivascular infiltration of immune cells, and signs of thrombosis (Dumkova et al., 2017).
Exposure to a higher concertation of PbO (2.23 x 106 particles/cm3) for 3 days, 2, 6, and 11 weeks led to
BLLs ranging from 10.4 (ig/dL at 2 weeks up to 17.4 (ig/dL after 11 weeks of exposure. The BLL at
3 days was not reported. Histological signs of cellular infiltration and alveolar septal wall thickening was
observed after 6 and 11 weeks of PbO exposure along with signs of macrophage proliferation (PCNA-
staining) (Dumkova et al., 2020b). These effects were not reported for the two-week exposure or an acute
3-day exposure to PbO. Despite increased signs of lung inflammation, signs of fibrosis and apoptosis
were not observed. Interestingly, a 5-week recovery period with no PbO exposure following 6 weeks of
PbO exposure was able to reduce both the lung Pb concertation and partially recover the histopathological
signs of inflammation seen at 6 weeks of PbO (Dumkova et al., 2020b).
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In a separate experiment, a similar procedure as Dumkova et al. (2020b) was followed using more
soluble Pb(N03)2nanoparticles in place of PbO. Mice were exposed to Pb(NC>3)2 particles for either
3 days, 2 weeks, 6 weeks, or 11 weeks and a separate recovery group that was exposed to Pb(NC>3)2 for
6 weeks and then filtered air for 5 weeks (Dumkova et al.. 2020a). Similar to the results with PbO,
Pb(NC>3)2 exposure showed an increase in histological signs of inflammation and lung damage.
Histological effects with Pb(NC>3)2 particle exposure were seen starting at 2 weeks of exposure and did
not completely resolve in the recovery group. Exposure to Pb(NC>3)2 reduced the number of lung
macrophages (CD68 positive stained cells) in the lung tissue which corresponded to an increase in
neutrophils (Myeloperoxidase positive cells) and mastocytes (Toluidine blue staining). Similar to the
findings with PbO, a 5-week recovery period with no Pb(N03)2 exposure following 6 weeks of Pb(N03)2
exposure was able to reduce both the lung Pb concertation and partially recover the histopathological
signs of inflammation. While macrophage number was partially restored after a 5-week recovery period,
the level of mastocytes remained elevated. Lung mRNA for inflammatory genes like IL-1B, IL-la, and
tumor necrosis factor-a were largely unchanged however RNA levels of NF-kB and IL6 were suppressed
after 3 days and 11 weeks of Pb(N03)2 suggesting that Pb(N03)2 dysregulates the inflammatory response
in the lung. While the data presented in these studies are mostly qualitative, it provides some preliminary
evidence of respiratory effects from inhalation of either Pb(N03)2 or PbO nanoparticles.
9.7.5 Summary and Causality Determination
The effects of Pb on asthma incidence and host defense, which includes data related to host
response to lung infection, are analyzed in the context of allergic disease and immune suppression
(Section 6.7.1 and Section 6.7.2).
The 2013 Pb ISA determined that the evidence for respiratory effects was "inadequate to
determine a causal relationship between Pb exposure and respiratory effects in populations without
asthma." This determination was based on inconsistent findings among studies and the limited quantity
and quality of both epidemiologic and experimental toxicological evidence of respiratory effects. While
there was some epidemiologic evidence of an association between short-term increases in ambient air Pb
and decreases in lung function, these studies were not informative to the causality determination due to
notable uncertainties regarding the size distribution of ambient air Pb, the relationship between ambient
air Pb and BLLs, and the confounding effects of co-occurring pollutants.
Evidence evaluated in the 2013 Pb ISA showed inconsistent relationships between BLLs and
bronchial responsiveness and lung function. Results from recent epidemiologic studies of the effect of
blood Pb on lung function and respiratory symptoms in children remain inconsistent (Section 9.7.3.1). In
adults, a new prospective cohort study provides evidence of accelerated lung function decline in those
with higher BLLs (Pak et al.. 2012). however the relationship between lung function decrements and
BLLs is inconsistent in a few recent cross-sectional analyses (Section 9.7.3.2). This lack of consistency in
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the epidemiologic literature is compounded by uncertainty related to exposure assessment and relative
lack of adjustment for correlated air pollutants. Toxicological data in the 2013 Pb ISA was mostly
limited to studies of concentrated ambient PM of which Pb was a component within a mixture of
pollutants, leaving uncertainty for the role of Pb in the observed effects. New toxicological studies
evaluating inhalation of Pb particles are limited in number but do provide evidence of gross histologic
signs of transient inflammation and lung damage; however, these data are largely qualitative and the
impact of these changes on lung function are unknown. Uncertainty remains regarding the relative size
distribution of Pb particles in ambient air and thus how well experimental generation of Pb particles
reflects ambient concentrations and particle size distribution.
Given the lack of consistency across a small body of epidemiologic evidence and uncertainty in
the direct relevance of a limited number of toxicological results to human lung function, the evidence is
not sufficient to draw a conclusion regarding causality. Thus, the cumulative body of evidence is
inadequate to infer the presence or absence of a causal relationship between Pb exposure and
respiratory effects in populations without asthma.
9.8 Mortality
9.8.1 Introduction, Summary of the 2013 Pb ISA, and Scope of the
Current Review
In the 2013 Pb ISA (U.S. EPA. 2013). the strongest evidence for Pb-associated mortality was
from studies examining cardiovascular mortality. The evidence did not provide strong support for Pb-
associated mortality other than through cardiovascular pathways, and very few studies examined total
(nonaccidental) mortality. For these reasons, the 2013 Pb ISA evaluated studies of all-cause mortality
together with studies examining cardiovascular mortality, and these studies were all included within the
cardiovascular disease chapter. Although this evidence contributed to the "causal relationship" between
Pb exposure and coronary heart disease, there were no distinct causality determinations for total or cause-
specific mortality. In the 2024 Pb ISA, the strongest evidence for Pb-associated cause-specific mortality
continues to come from studies of cardiovascular mortality. However, additional studies examining total
non-accidental mortality have become available since the 2013 Pb ISA, and this section discusses and
evaluates those studies. Studies that examine cardiovascular-related mortality or other cause-specific
mortality are discussed in detail within the appropriate outcome-specific appendices (e.g., cardiovascular
disease [CVD]-related mortality is discussed in Appendix 4) and are briefly summarized in this section.
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9.8.2
Scope
The scope of this section is defined by PECOS statements. The PECOS statement defines the
objectives of the review and establishes study inclusion criteria thereby facilitating identification of the
most relevant literature to inform the Pb ISA.37 In order to identify the most relevant literature, the body
of evidence from the 2013 Pb ISA was considered in the development of the PECOS statements for this
Appendix. Specifically, well-established areas of research; gaps in the literature; and inherent
uncertainties in specific populations, exposure metrics, comparison groups, and study designs identified
in the 2013 Pb ISA inform the scope of this Appendix. The 2013 Pb ISA used different inclusion criteria
than the 2024 Pb ISA, and the studies referenced therein often do not meet the current PECOS criteria
(e.g., due to higher or unreported biomarker levels). Studies included in the 2013 Pb ISA, including many
that do not meet the current PECOS criteria, are discussed in this appendix to establish the state of the
evidence prior to this assessment. Except for supporting evidence used to demonstrate the biological
plausibility of Pb-associated effects on mortality, recent studies were only included if they satisfied all the
components of the following PECOS statements:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect.
Exposure: Exposure to Pb38 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure39; or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles).
Outcome: Mortality.
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
37The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
38Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area of particular
relevance to the NAAQS review (e.g., longitudinal studies designed to examine recent versus historical Pb
exposure).
39Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 (im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 |im3 (PM2.5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is difficult
to assess the representativeness of these concentrations to population exposure [Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PNfc.s with BLLs are lacking.
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endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
9.8.3 Total (non-Accidental) Mortality
The 2013 Pb ISA (U.S. EPA. 2013) evaluated a small number of studies that examined the
association between biomarkers of Pb exposure and all-cause mortality. Overall, these studies reported
consistently positive associations between Pb biomarkers and all-cause mortality. Specifically, Lustberg
and Silbergeld (2002) indicated an increased risk of all-cause mortality when comparing the highest
tertiles of BLLs (20-29 (ig/dL) to the lowest (<10 (ig/dL). Lustberg and Silbergeld (2002) conducted this
analysis among NHANES II cohort, which had high BLLs (mean 14 (.ig/dL). Additionally, Schober et al.
(2006) and Menke et al. (2006) both evaluated the NHANES III cohort, which had an overall lower BLL
(mean: 2.6 |ig/dL). and still identified a positive association between BLLs and all-cause mortality
(Figure 9-1). Notably, both NHANES cohorts included adult study populations with higher past than
recent Pb exposures, making it difficult to characterize the specific timing, duration, frequency, and level
of Pb exposure that contributed to the observed associations. Recent evidence continues to support the
association between Pb biomarkers and all-cause mortality. Study-specific details, including biomarker
Pb levels, study population characteristics, confounders, and select results from these studies, are
highlighted in Figure 9-3 and Table 9-17. Studies in Figure 9-3 are standardized to be interpreted as the
risk of all-cause mortality associated with a 1 (ig/dL increase in BLL. Study details in Table 9-10 include
standardized results as well as results that could not be standardized based on the information provided in
each paper. An overview of the recent evidence is provided below.
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Reference Study Population Pb distribution
Menke et al, 2006 NHANES III Adults £ 20 Mean: 2.58
Pb measurement -Years of
year follow-up
Lanphearetal,2018 NHANES III Adults a 20
Schoberetal,2006 NHANES III Adults £40
Geometric Mean: 2.71
Geometric SE: 1.31
Median
T1 (2.6)
T2 (6.3)
T3 (11.8)
Median
van Bemmeletal, 2011 NHANES III Adults £40 <5ug/dL2.6
a 5 ug/dL7.5
Duan et al, 2020*
NHANES Adults 2 20
Median (IQR)
1.49(0.93,2.31)
1988-1994
1988-1994
1988-1994
1988-1994
1999-2014
12
19
8.6
7.1
all cause
all cause
all cause
7.5-7.8 all cause
all cause ALAD GG
all cause ALAD CG/GG
all cause
I 1 1
0.90 1.00 1.20
Hazard Ratio (95% CI) per 1 ug/dL increase in blood Pb
1.40
ALAD GG and ALAD CG/GG = variants of 5-aminolevulinic acid dehydratase, T# = fertile #, NHANES = National Health and Nutrition Examination Survey.
Note: Red text: Studies published since the 2013 Pb ISA; Black text: Studies included in the 2013 Pb ISA.
Effect estimates are standardized to a 1 |jg/dL increase in blood Pb. If the Pb biomarker is log-transformed, effect estimates are standardized to the specified unit increase for the
10th—90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated interval.
*Study estimated relative risk.
Figure 9-3 Effect estimates for associations of blood Pb with all-cause mortality.
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In a recent extended analysis of the NHANES III cohort, Lanphear et al. (2018) increased the
average follow-up time of the Menke et al. (2006) analysis by over 7 years (from 12 to -19 years),
resulting in a substantial increase in the number of total deaths observed (4,222 versus 1,661). Lanphear
et al. (2018) reported that 1 (ig/dL higher BLLs were associated with a hazard ratio (HR) of 1.06 [95%
CI: 1.03, 1.09]) for all-cause mortality. The authors also calculated the population attributable fraction for
both all-cause and cardiovascular mortality, to estimate the proportional reduction in mortality that would
be expected if BLLs in those >20 were reduced to 1 (ig/dL. Lanphear et al. (2018) estimated that the
population attributable fraction for all-cause mortality was 18% (95% CI: 10.9, 26.1%), while the
population attributable fraction for cardiovascular mortality was 28.7% (95% CI: 15.5, 39.5%). Therefore,
given the proportion of all-cause mortality attributable to cardiovascular causes (both in this study [~38%]
and nationally [-33%; NHLBI, 2017, 3980932}]), CVD mortality likely accounts for a large proportion
but not the entirety of the all-cause mortality signal. The authors also used a five-knot restricted cubic
spline analysis to assess potential non-linearities and observed a generally sigmoidal concentration-
response (C-R) relationship between BLLs and all-cause mortality, with some attenuation of the C-R
relationship below 2.5 (ig/dL (Figure 9-4). The general shape of the C-R relationship is consistent with
previous results from Menke et al. (2006).
4~\
0 2.5 5 7.5 10
Concentrations of lead in blood |ig/dL
Note: Restricted cubic spline (5 knots) (red line) and adjusted HRs (black line) with 95% CI's (hatched lines) for all-cause mortality.
Source: Adapted from Lanphear et al. (2018).
Figure 9-4 Dose-response relationship between blood Pb levels and
all-cause mortality.
Other recent studies also evaluated the relationship between blood Pb and total mortality using
NHANES data. Using NHANES III, van Bemmel et al. (2011) estimated an increased association
between BLLs and all-cause mortality (HR: 1.04 [95% CI: 0.98, 1.10]). In addition, van Bemmel et al.
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(2011) also evaluated this relationship by polymorphisms in 5-aminolevulinic acid dehydratase (ALAD).
A critical mechanism of Pb toxicity is its ability to interact and inhibit key enzymes, such as ALAD, in
the heme biosynthesis pathway. This study evaluated associations between BLLs, and mortality stratified
by ALAD variant (ALADGG [more common genotype] or ALADCG/GG). However, there was little
difference between the estimates generated when stratified (ALADGG HR: 1.03 [95% CI:0.98, 1.08],
ALADCG/GG HR: 1.09 [95% €1:0.93. 1.28]), when comparing BLLs >5 (ig/dL to levels <5 (ig/dL.
Using more recent NHANES cycles (1999-2014), Duan et al. (2020) also reported a positive association
between blood Pb and all-cause mortality (RR: 1.39 [95% CI: 128, 1.51]). In a similar analysis using
recent KNHANES cycles (2007-2015), Bvun et al. (2020) evaluated the association between BLLs and
total (nonaccidental) mortality using KNHANES (2007-2015) baseline data, and mortality data linked
through 2018. Overall, there were positive associations between higher tertiles of blood Pb exposure and
all-cause mortality. Compared to the first tertile of BLLs (<1.91 |ig/dL). the HR for all-cause mortality
was 2.02 (95% CI: 1.20, 3.40) for the second tertile (1.91-2.71 ^ig/dL) and 1.91 (95% CI: 1.13, 3.23) for
the third tertile (>2.71 (.ig/dL).
In addition to studies using nationally representative survey data, a recent study by Hollingsworth
and Rudik (2021) implemented a quasi-experimental design to examine the effect of the phase out of
leaded gasoline in automotive racing on mortality rates in older adults. Comparing time periods prior to
and after the phaseout of leaded gasoline in professional racing series (i.e., the National Association for
Stock Car Auto Racing [NASCAR] and the Automobile Racing Club of America [ARCA]), the authors
used a difference-in-differences technique to estimate county-level changes in air Pb concentrations,
elevated BLL prevalence among children, and mortality rates in race counties and counties bordering race
counties relative to control counties. A detailed discussion of results for air Pb concentrations and BLLs is
presented in Section 2.4.1. In short, there were substantial declines in both air Pb concentrations and the
prevalence of children with elevated BLLs associated with the phaseout of leaded gasoline. The authors
also reported significant declines in mortality rates over this same period. Specifically, in the period
following de-leading of gasoline, there was an estimated decline in annual age-standardized all-cause
mortality rates of 91 deaths per 100,000 in race counties and 38 deaths per 100,000 in border counties.
Similar to the exposure results, the mortality estimates appear to demonstrate a distance gradient.
Although this analysis includes county-level data, the difference-in-difference approach controls for
spatially varying confounders by estimating the difference in mortality rates in adjoining years in the
same county and controls for temporally varying confounders by assessing the difference of those
differences between locations. The authors additionally adjust for potential confounders that may vary
spatially and temporally (e.g., unemployment rate and quantity of Toxic Release Inventory [TRI] Pb
emissions). Hollingsworth and Rudik (2021) did not adjust for potential co-pollutant exposures, but
provide evidence that there is no differential effect of leaded and unleaded races on other co-pollutant
concentrations (i.e., CO, VOCs, PMio, PM2.5, NO2, and O3) in the weeks leading up to and following the
race. However, because the mortality rates are an annual measure, there is remaining uncertainty
regarding potential differential trends in the long-term average of other pollutants that could be correlated
with the phaseout of leaded gasoline in NASCAR and ARCA.A recent re-analysis of NAS data
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(Wcisskopf et al.. 2015). expanded on a similar analysis (Wcisskopf et al.. 2009) that was discussed in the
2013 Pb ISA. In the re-analysis, special considerations for selection bias were taken to account for the
probability that older individuals who elected to participate in the study were more likely to be free of
cardiovascular disease than those who declined to participate. Specifically, the authors created four
different models, which controlled for different covariates, additional markers for SES, and restricted by
age. In this analysis, the authors restricted the sample (Model 3 and Model 4) to participants that were
<45 years at the start of the NAS study, since cardiovascular disease-related deaths would be relatively
rare in the younger population and would therefore not impact study participation. This study indicated a
positive association with all-cause mortality (HR: 1.86 [95% CI: 1.12, 3.09]) when comparing the highest
tertile (>31 |ig/g) of patella Pb to the lowest tertile (<20 |ig/g). in the model restricting the age of
participants to participants <45 years at the start of the NAS study. No associations were observed
without the age restriction or with blood or tibia Pb.
Since Pb has been identified as being associated with renal insufficiency, previous studies have
further assessed if Pb accumulates in patients with end-stage renal disease (ERSD). In a recent
prospective cohort study in Taiwan, Lin et al. (2011) followed study subjects on maintenance
hemodialysis for a period of 18 months. Overall, subjects included in the study had higher BLLs (mean:
11.5 (ig/dL) than the general Taiwanese population (mean: 7.7 (ig/dL). It is suspected that hemodialysis
patients may experience higher BLLs since their kidneys may no longer be able to excrete Pb from the
body due to a total loss of renal function (Appendix 5). Among this group, there was a strong but
imprecise association between BLLs and all-cause mortality when comparing those in the second tertile
of BLLs (8.51-12.64 (ig/dL) to those in the first tertile of BLLs (<8.51 (ig/dL; HR: 2.69 [95% CI: 0.47,
3.44]). This effect was higher in magnitude, but even more imprecise among those in the third tertile of
BLLs (>12.64 (ig/dL) compared with the first tertile (HR: 4.70 [95% CI: 1.92, 11.49]). The imprecise
effect estimates in this analysis are likely due to a combination of the relatively small sample size and
short follow-up period, leading to a small number of deaths included in the analysis. The small number of
cases reduces statistical power, as well as the likelihood that an observed result reflects a true effect.
In addition, several analyses evaluated metal chelation therapy as a treatment for those with
atherosclerotic plaques and evaluated subsequent all-cause mortality outcomes in the Trial to Assess
Chelation Therapy (TACT) study. The TACT study was a randomized control trial (RCT) with a 2 x 2
factorial design evaluating chelation therapy with ethylenediaminetetraacetic acid (EDTA) plus the use of
high dose oral vitamins. The factorial group results indicated that a combination of EDTA and high-dose
vitamins was associated with a reduction in deaths from all causes (Lamas et al.. 2014). In the same trial,
the findings indicated that diabetic patients >50 years had a reduction (10% versus 16% HR: 0.59 [95%
CI: 0.44, 0.79]) in the number of deaths from all-causes following EDTA chelation therapy (Escolar et al..
2014). Although these studies suggest a clear association between chelation therapy and a reduction in
overall deaths, it should be noted that most of these studies did not measure BLL pre and post chelation.
Notably, chelation therapy reduces levels of other heavy metals in the blood and thus does not establish a
direct effect of Pb reduction absent direct measures of metal biomarkers. Thus, chelation therapy in
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populations with low BLLs is an area of research that could be expanded to potentially provide strong
quasi-experimental support for other lines of evidence that quantitatively describe the associations
between Pb biomarkers and all-cause mortality, as well as other health effects. A follow-up RCT,
TACT2, a replicative study in diabetics with a history of MI, is currently underway to confirm the results
reported as a result of TACT (Lamas and Ergui. 2016).
In contrast to the generally consistent evidence of an association between BLLs and all-cause
mortality, a small Canadian study evaluating several trace metals observed a null association between all-
cause mortality and BLLs among hemodialysis patients (>18 years of age) (Tonelli et al.. 2018). Patients
in this cohort had relatively low BLLs (1st decile: 0.06 (ig/dL, 10th decile 1.74 (.ig/dL). and there was no
observed relationship between BLLs and all-cause mortality when comparing the highest to the lowest
decile. The authors only presented quantitative results for statistically significant associations, so it is
unclear whether there was any evidence of a non-statistically significant association. Additionally, Tonelli
et al. (2018) was likely underpowered to detect a HR in the range reported in other studies of BLLs and
all-cause mortality (Figure 9-4).
9.8.4 Cause-Specific Mortality
The mortality studies available for review in the 2013 Pb ISA focused primarily on
cardiovascular mortality, and consistently reported positive associations with overall cardiovascular
mortality and cause-specific cardiovascular mortality. Recent studies also evaluate cardiovascular
mortality in addition to other cause-specific mortality outcomes.
Recent analyses further indicate a positive association between Pb exposure and cardiovascular
mortality and are further described in Section 4.10. In summary, there were several studies using
nationally representative data with low BLLs (mean <2 (ig/dL) that consistently reported increased
associations between biomarkers of Pb exposure and cardiovascular mortality. However, these
populations were largely similar (mostly from NHANES III or other more recent NHANES cycles) and
still include individuals with sizeable historic exposures to Pb. For specific causes of CVD mortality
(e.g., myocardial infarction (MI), ischemic heart disease (IHD), stroke), the measures of association were
higher in magnitude but were less precise (i.e., wider 95% CIs), likely due to the smaller number of
cause-specific cardiovascular-related deaths. Additionally, in the quasi-experimental study discussed in
Section 9.8.3, deleading of racing gasoline led to declines in county-level cardiovascular mortality rates
(Hollingsworth and Rudik. 2021). Evidence from RCT trials evaluating chelation therapy (Escolar et al..
2014; Lamas et al.. 2014) indicates reductions in cardiovascular mortality following chelation with EDTA
and high doses of oral vitamins, yet the study did not specifically evaluate BLLs before or after chelation
therapy. This evidence helps to strengthen the evidence base indicating an association between
biomarkers of Pb exposure and increased risk of cardiovascular mortality.
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Several recent studies also evaluated the relationship between Pb exposure biomarkers and cancer
mortality, as described in Section 10.4. In summary, there were a limited number of studies evaluating Pb
biomarkers of exposure and overall cancer mortality. Most studies relied on nationally representative data
and yielded inconsistent but mostly null associations between Pb exposure and cancer mortality.
However, the follow-up period in many of these analyses was short (<11 years), with a small number of
cancer deaths and a lack of control of some potential influential confounders, such as comorbidities and
BMI.
Additionally, some studies evaluated alternative cause-specific mortality outcomes. A cohort
study analyzed data from five NHANES cycles (1999-2008) and reported a positive, but imprecise
association between blood Pb and Alzheimer's disease (AD) mortality rSection 3.5.4; (Horton et al..
2019)1. The imprecise effect estimate is likely due to the small number of AD mortality cases (n = 81)
that resulted from AD mortality being determined by the listing of the immediate cause of death rather
than the underlying cause of death. Additionally, Lin et al. (2011) prospectively evaluated subjects on
maintenance hemodialysis for a period of 18 months and evaluated infection-caused mortality. Among
this group there was an imprecise increase in mortality (HR: 5.35 [95% CI: 1.38, 20.83]) in the highest
tertile (>12.64 (ig/dL) compared to the lowest tertile (<8.51 (ig/dL). This association persisted (HR: 4.72
[95% CI: 1.27, 17.54]) even after correction for hemoglobin (dividing BLL by hemoglobin
concentration). Finally, a quasi-experimental reported a decrease in county-level respiratory mortality
rates in association with the phase out of leaded gasoline in automotive racing (Hollingsworth and Rudik.
2021).
9.8.5 Biological Plausibility
In evaluating the biological plausibility of reported associations between Pb exposure and total
non-accidental mortality, this section considers the biological evidence supporting health outcomes likely
to contribute to total mortality. As summarized above, studies consistently report positive associations
between Pb exposures and cardiovascular-related mortality, with much more limited evidence for
associations with other causes of mortality. Overall, cardiovascular mortality is the most common
contributor to total non-accidental mortality (i.e., accounting for about 33% of total mortality) (NHLBI.
2017). As it pertains to Pb exposure, the available evidence provides strong support for Pb-associated
cardiovascular effects and supports a continuum of effects leading to cardiovascular mortality, as
described further in Appendix 4. Direct evidence for cardiovascular effects following Pb exposures comes
from numerous animal toxicological studies, and there is coherence between these animal studies and
epidemiologic studies that report associations with some of the same cardiovascular outcomes
(e.g., increased blood pressure, changes in cardiac electrophysiology). Animal studies additionally
support the biological plausibility of the consistent epidemiologic associations reported between body Pb
concentrations and cardiovascular outcomes such as hypertension and cardiovascular mortality.
Section 4.10 characterizes the strong evidence indicating the mechanisms by which exposure to Pb could
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plausibly progress from initial events to endpoints relevant to the cardiovascular system, such as
hypertension, exacerbation of IHD, and potential MI or stroke. In particular, exposures to Pb can result in
oxidative stress and systemic inflammation, which could potentially lead to impaired vascular function, a
pro-atherosclerotic environment, and increases in blood pressure. There is animal toxicological evidence
demonstrating all of these effects following exposure to Pb (Section 4.8). Atherosclerosis and increased
blood pressure can then set the stage for an MI or stroke that could result in mortality. Thus, the
progression demonstrated in the available evidence for cardiovascular morbidity supports potential
biological pathways by which Pb exposure could result in cardiovascular mortality.
The current evidence strongly supports a plausible relationship between Pb exposure and
cardiovascular mortality. Additionally, Pb may act on other biological pathways leading to death. There is
some limited evidence that BLLs are associated with other causes of mortality, including AD and
infection. The strongest evidence for biologically supported pathways leading to neurodegenerative
disease include the effect of Pb on cellular protein function and subsequent initiation of oxidative stress-
and inflammation-mediated pathways (Section 3.3). AD, specifically, has been linked with increased
markers of neuroinflammation. Studies with exposure of postweaning animals to Pb have shown
increased inflammation associated with AD markers, as well as inhibition of AD markers following
postexposure treatment with anti-inflammatory and antioxidative molecules. Regarding infection-related
mortality, biological plausibility for the observed association is provided by toxicological and
epidemiologic studies demonstrating (1) skewing of T cell populations, promoting Th2 cell formation and
cytokine production, (2) decreased IFN-y production, (3) decrements in macrophage function, (4)
production of inflammatory mediators, and (5) disruption of the microbiome, all of which could lead to
immunosuppression (Section 6.6.1).
9.8.6 Summary and Causality Determination
The 2013 Pb ISA did not make a causality determination regarding the relationship between Pb
exposure and total (nonaccidental) mortality, but studies examining this relationship did support the
coronary heart disease causality determinations made within the cardiovascular disease chapter. The
evidence available at the time of the last review was limited, but reported consistently positive
associations between Pb biomarkers and all-cause mortality (Menke et al.. 2006; Schober et al.. 2006).
These results were additionally supported by consistent positive associations between BLLs and overall
cardiovascular mortality (Section 4.10) as well as cause-specific cardiovascular mortality (e.g., MI, IHD,
stroke)). Menke et al. (2006) examined the shape of the C-R relationship between BLLs and all-cause
mortality using quadratic spline models, which generally appeared to support a linear, no-threshold
relationship, although the HRs were somewhat attenuated at BLLs <2.5 (ig/dL. Notably, the majority of
mortality studies analyzed participants from NHANES cohorts, either NHANES II or NHANES III, so
while the results are consistent, they do not represent entirely independent study populations.
Additionally, while some of the studies evaluated in the 2013 Pb ISA examined populations with low
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mean BLLs (<3 (.ig/dL). study participants were born prior to the phase-out of leaded gasoline and
therefore likely had much higher past Pb exposures, making it difficult to characterize the specific timing,
duration, frequency, and level of Pb exposure that contributed to the observed associations.
Prospective cohort studies evaluated since the completion of the 2013 Pb ISA continue to provide
consistent evidence of positive associations between Pb exposure and total (nonaccidental) mortality.
Many recent analyses further evaluated the association between BLLs and the risk of mortality using
NHANES cohorts linked to mortality databases, including an extended analysis of the NHANES III
cohort with additional years of follow-up (Lanphear et al.. 2018) and analyses of more recent NHANES
cycles (Bvun et al.. 2020; Duan et al.. 2020; van Bemmel et al.. 2011). In addition to NHANES analyses,
another analysis of participants from a nationally representative survey [KNHANES; (Bvun et al.. 2020)1
and a smaller prospective cohort study of hemodialysis patients (Lin et al.. 2011) provide evidence of an
association between BLLs and total (non-accidental) mortality. These findings are supported by a quasi-
experimental study that reported a decline in county-level all-cause mortality rates following the phase
out of leaded gasoline in automotive racing (Hollingsworth and Rudik. 2021). Recent studies continue to
include populations with low mean blood Pb concentrations, but do not address potentially large
differences in past versus current exposures. Thus, there is remaining uncertainty as to the specific timing,
duration, frequency, and level of Pb exposure that contributed to the observed associations. The observed
associations between BLLs and total mortality are large in magnitude (Figure 9-3), though uncertainty in
the levels of Pb exposure that contributed to the observed associations may also introduce uncertainty in
estimating the magnitude of the effect. One recent study examined the C-R relationship between blood Pb
and total mortality (Lanphear et al.. 2018). Similar to Menke et al. (2006). Lanphear et al. (2018)
observed generally sigmoidal spline curves with some evidence of attenuation of the C-R relationship
below 2.5 (ig/dL (Figure 9-4).
The body of evidence for total mortality is further supported by strong evidence for a causal
relationship between Pb exposures and cardiovascular effects and cardiovascular mortality (Section 4.10).
Cardiovascular mortality comprises a large portion of total mortality (Section 4.10), and recent studies
consistently report positive associations with BLLs. The recent evidence includes a wider range of study
populations and expanded evidence on the C-R relationship that generally supports a linear relationship
between BLLs and cardiovascular mortality, with no evidence of a threshold. There is also coherence of
effects across the scientific disciplines (i.e., animal toxicological, controlled human exposure, and
epidemiologic studies) and biological plausibility for Pb-related cardiovascular disease (Appendix 4).
which provides additional support for the Pb-mortality relationship.
Overall, there is sufficient evidence to conclude that there is a causal relationship between Pb
exposure and total (nonaccidental) mortality. This conclusion is driven by epidemiologic evidence for
Pb-associated all-cause mortality and the strong epidemiologic and experimental animal evidence
supporting a causal relationship with cardiovascular effects and cardiovascular mortality. Recent
epidemiologic studies build upon evidence from the 2013 Pb ISA and provide largely consistent evidence
9-71
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of an association between biomarkers of Pb exposure and total mortality. A few uncertainties remain in
the evidence base, including a limited number of studies and analyses of similar or overlapping study
populations. However, these studies are supported by more robust evidence of Pb-related cardiovascular
mortality, which comprises nearly 33% of total mortality. In addition, evidence for cardiovascular
morbidity provides biologically plausible pathways through which Pb exposure could result in mortality.
There is also very limited evidence that Pb exposure is positively associated with other causes of
mortality, including AD and infection. Biological plausibility for these outcomes is demonstrated by
pathways leading from Pb exposure to neurodegenerative disease and immunosuppression, respectively.
However, although there is toxicological evidence that developmental exposure to Pb increases the
expression of proteins related to AD, the epidemiologic evidence relating Pb exposure to incident AD
remains limited. The key evidence, as it relates to the causal framework, is summarized in Table 9-3.
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Table 9-3 Summary of evidence for a causal relationship between Pb exposure and total mortality
Rationale for Causality
Determination3
Key Evidence"
Key References"
Pb Biomarker Levels Associated
with Effects0
Consistent epidemiologic
evidence from multiple studies
at relevant BLLs
Increases in total mortality in multiple nationally
represented studies. Total mortality
associations are further supported by increases
in cardiovascular mortality conducted within
nationally represented studies.
(Hollinasworth and Rudik. 2021: Bvun et
al., 2020: Duan et al.. 2020: Lanphear et
al., 2018: van Bemmel et al., 2011: Menke
et al., 2006)
Median, Mean, and Geometric
Mean BLLs: 1.49-2.71 pg/dL
Epidemiologic evidence
supports no evidence of a
threshold between Pb
biomarkers of exposure and
total mortality at the
concentration ranges
examined
Recent studies provide direct evidence of a
linear or sigmoidal, no-threshold C-R
relationship at lower concentrations of BLLs.
(Menke etal.. 2006)
(Lanphear et al.. 2018)
Mean BLL: 2.58 pg/dL
Geometric Mean BLL: 2.71 pg/dL
Biological plausibility from
cardiovascular morbidity
evidence
Stronq evidence for coherence of effects across Appendix 4
scientific disciplines and evidence for a range of
cardiovascular effects in response to increases
in biomarkers of Pb exposure, especially for
increases in blood pressure and hypertension.
The collective body of cardiovascular morbidity
evidence provides biological plausibility for a
relationship between biomarkers of Pb
exposure and cardiovascular mortality, which
comprises -33% of total mortality.
BLLs = blood lead levels; C-R = concentration-response; Pb = lead.
aBased on aspects considered in judgments of causality and weight-of-evidence in causal framework
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
causality determination and, where applicable, to uncertainties or
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9.9
Evidence Inventories - Data Tables to Summarize Study Details
Table 9-4
Epidemiologic studies of exposure to Pb and hepatic effects
Reference and Study
Design
Study Population
Exposure Assessment Outcome
Confounders
Effect Estimates and 95%
Clsa
Direct Evaluation of Liver Injury
tZhai et al. (2017)
Yangtze River Delta
Region
China
1 yr (2014)
Cross-sectional
SPECT-China
n = 2011
General population,
>18 yr old with no history
of excessive alcohol
consumption or viral
hepatitis
Blood
Pb measured in venous
whole blood using
atomic absorption
spectrometry
Age at measurement:
>18 yr old
Median:
Males: 5.29 |jg/dL
Females: 4.49 |jg/dL
25th:
Males: 3.61 |jg/dL
Females: 2.98 |jg/dL
75th:
Males: 7.28 |jg/dL
Females: 6.59 |jg/dL
Nonalcoholic fatty liver
disease
Two doctors
performed abdominal
ultrasounds and
categorized liver
status as normal or
fatty using predefined
criteria
Age at outcome:
>18 yr old
Age, region,
education, current
smoking, current
drinking, ALT,
diabetes, waist
circumference, BMI,
LDL cholesterol, HDL
cholesterol,
triglycerides, total
cholesterol, and
blood cadmium levels
ORs for NAFLD prevalence
across blood Pb quartiles
Males
Q1:
Ref.
Q2:
1.70 (0.84,
3.42)
Q3:
1.84 (0.88,
3.86)
Q4:
2.17 (0.99,
4.75)
Females
Q1:
Ref.
Q2:
1.38 (0.96,
2.00)
Q3:
1.50 (1.02,
2.18)
Q4:
1.61 (1.08,
2.41)
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tWerder et al. (2020) Gulf Long-Term Follow-up Blood
Gulf Region
United States
2012-2013
Cross-sectional
Study
n = 214
Non-smoking >30 yr old
male oil spill response
workers and oil spill safety
trainees with no history of
liver disease or heavy
alcohol use
Pb measured in venous
whole blood using solid-
phase micro-extraction
with gas
chromatography/mass
spectrometry
Age at measurement:
>30
Liver injury
Cytokeratin 18 (CK18
M65 and CK18 M30)
Age at outcome:
>30
Age, race, alcohol
consumption, serum
cotinine, BMI,
diabetes dx, and
education
Change in CK18 M65 (U/L)
2.4 (-12.69, 17.49)
Change in CK18 M30 (U/L)
21.7 (9.94, 33.46)
Mean: 1.82 (1.76)
tChuna et al. (2020)
South Korea
2 yr (2016-2017)
Cross-sectional
KNHANES
n = 4420
Adults, >20 yr old
Blood
Pb measured in venous
whole blood using
GFAAS
Age at measurement:
>20 yr old
Mean: 1.81 pg/dL
Max: 20.16 pg/dL
Hepatic steatosis and
fibrosis
Hepatic steatosis (HS)
as indicated by an HS
Index = 36
(8 x (ALT/AST ratio) +
BMI (+2 if female; +2
if had diabetes
mellitus)). Hepatic
Fibrosis (HF) as
indicated by a fibrosis-
4 (FIB-4) score >2.67
((age * AST
level)/(platelet level *
v(ALT level)).
Age at outcome:
>20 yr old
Age, smoking status,
alcohol consumption,
hypertension status,
obesity status,
diabetes status,
hypertriglyceridemia
status, blood Hg,
blood Cd.
ORs
Hepatic Steatosis
Men
0.83 (0.66, 1.03)
Women
0.98 (0.80, 1.19)
Fibrosis
Men
0.70 (0.44, 1.09)
Women
0.72 (0.42, 1.26)
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tReia et al. (2020)
United States
5 yr (2011-2016)
Cross-sectional
NHANES
n = 2499
General population >20 yr
old with nonalcoholic fatty
liver disease (NAFLD)
Blood
>20 yr old
Mean: 1.01 pg/dL
75th: 1.62 pg/dL
Liver fibrosis
NAFLD Fibrosis Score
Age at outcome:
>20 yr old (concurrent
with exposure)
Age, gender, waist
circumference,
hypertension, liver
function test,
hemoglobin A1c,
triglycerides,
smoking, and PIR
ORs (NAFLD Fibrosis Score
>0.676)
Q1
Q2
Q3
Q4
Reference
2.79 (1.39, 5.63)
3.74 (2.01, 6.96)
5.93 (2.88, 12.24)
Serum Biomarkers of Liver Function
tPollack et al. (2015)
BioCycle
Blood
ALT, ALP, AST,
Linear mixed models
AST (% change):
n =259
Bilirubin
adjusted for age,
5.02 (-1.36, 11.41)
Buffalo, NY
Pb measured in venous
BMI, race, average
ALT (% change):
United States
Premenopausal women
whole blood using ICP-
ALT (U/L), ALP (U/L),
calories, alcohol
2 menstrual cycles (8
followed for 2 menstrual
MS
AST (U/L), Bilirubin
intake, smoking, and
6.39 (3.07, 9.72)
visits per cycle) (2005-
cycles
(mg/dL)
cycle day
ALP (% change):
2007)
Age at measurement:
Cohort
27.4 (SD: 8.2)
Age at outcome:
2.14 (-5.05, 9.33)
27.4 (SD: 8.2)
1.03 pg/dL
tChen et al. (2019)
Guangdong
China
1 yr(2015)
Cross-sectional
n = 267
Hospitalized patients from
two regions in Guangdong
(one e-waste polluted
area and a matched
control area). Patients
with heart or kidney
disease, those taking
drugs with hepatic toxicity,
and those with a history of
alcohol consumption or
smoking were excluded.
Blood
Pb was measured in
venous whole blood
using GFAAS
Age at measurement:
4 to 85 yr old
Median:
Exposed: 8.7 pg/dL;
Control: 5.1 pg/dL
75th:
Exposed: 12.2 pg/dL;
Control: 8.4 pg/dL
Abnormal liver
function
Abnormal liver
function defined as
two transaminases
(AST, ALT, or GGT)
above normal range
or one at least two
times higher than
normal range (40 U/L)
Age at outcome:
4 to 85 yr old
(concurrent with
exposure)
Age, gender, hepatic
disease, RBC, Hb,
and platelets
OR for Prevalence of
Abnormal Liver Function
1.94 (1.00, 3.73)
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tChristensen et al.
(2013)
United States
2 yr (2003-2004)
Cross-sectional
NHANES
n = 1345
General population,
>12 yr old. No chronic
hepatitis or liver disease,
and no high alcohol
intake.
Blood
Pb measured in venous
whole blood using ICP-
MS
Age at measurement:
>12 yr old
Liver function
Serum ALT
Age at outcome:
>12 yr old
Sex, Race/Ethnicity,
Age, PIR, BMI
Change in ALT (U/L)
Q1
Q2
Q3
Q4
Reference
-0.068 (-0.14, 0.004)
-0.039 (-0.113, 0.035)
-0.103 (-0.185, -0.021)
Mean NR
tObena-Gvasi (2019) NHANES
United States
NHANES 2009-2016
Cross-sectional
n = 7,730 young adults
(18-44); 5,744 middle-
aged adults (45-65)
General population; ages
18-65
Blood
BLL measured in venous
whole blood using ICP-
MS
Age at measurement:
>18 yr old
Mean:
Young adults:
1.03 |jg/dL
Middle-aged adults:
1.62 |jg/dL
GGT (U/L)
Serum GGT (U/L)
Age at outcome:
>18 yr old
Gender, BMI,
income, ethnicity,
and alcohol
consumption
ORs (GGT >18 U/L)
Young Adults
1.94 (1.65, 2.28)
Middle-Aged Adults
1.34 (1.14, 1.58)
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Referen<^and Study stlldy Poplllatlon ExposlIre Assessment 0lItcome ConfolInders Effec, Es«i™,?s ,„d 95%
Serum Lipids
tPeters et al. (2012)
United States
Blood Pb measured
between 1999-2008;
Serum lipids measured
3 to 4 yr after blood Pb
Cohort
Normative Aging Study
n = 426
Older male Veterans
Blood, Bone
Blood Pb measured in
venous whole blood
using GFAAS
Serum lipids
Triglycerides, total
cholesterol, HDL-C,
LDL-C
Mean: 4.01 ± 2.30 |jg/dL Age at outcome:
3 to 4 yr after blood
Pb
Age at baseline, yr
between baseline
and outcome,
education, BMI,
alcohol intake,
smoking status,
pack-yr of smoking,
hypertension status,
and statin use
ORs
Total C (>200 mg/dL)
1.08 (0.99, 1.19)
LDL-C (>130 mg/dL):
1.02 (0.91, 1.15)
HDL-C (<40 mg/dL):
0.90 (0.80, 1.00)
Triglycerides (>200 mg/dL):
0.99 (0.87, 1.13)
tXu et al. (2021)
United States
NHANES 2005-2016
Cross-sectional
NHANES
n = 7457
General population; Ages
20 to 79 yr old
Blood
Pb measured in venous
whole blood samples
using ICP-MS
Age at measurement:
Mean (SD):
43.68 (15.02) yr
GM: 1.23 |jg/dL
Dyslipidemia
Total cholesterol,
LDL-C, non-HDL-C,
triglycerides
Age at outcome:
Mean (SD): 43.68
(15.02)
Age, sex, race, BMI,
education status,
smoking status,
alcohol consumption,
physical activity, PIR,
systolic blood
pressure, serum
cotinine, and Cd
RRs
Total C (>200 mg/dL)
1.01 (1.00, 1.01)
non-HDL-C (>160 mg/dL)
1.00 (0.99, 1.01)
LDL-C (>130 mg/dL)
1.02 (1.00, 1.04)
Triglycerides (>200 mg/dL)
0.99 (0.98, 1.00)
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tLee and Kim (2016) KNHANES
Korea
2005-2010
Cross-Sectional
n = 7559
Korean adults aged 20+
Blood
Pb measured in venous
whole blood using
GFAAS
Age at measurement
Mean (SD):
No MetS: 42.32 (0.294)
yr; MetS: 48.36 (0.574)
yr
Geometric Mean (SD)
No MetS: 2.73 (0.024)
|jg/dL; MetS: 2.96
(0.049) |jg/dL
Serum Lipids
Low HDL cholesterol
(<40 mg/dL in women
or <50 mg/dL in men);
Elevated serum
triglycerides
(=150 mmHg)
Age at outcome
same as age at
exposure assessment
Age, BMI, residence ORs
area, education level,
smoking and drinking
status, exercise,
serum aspartate
aminotransferase,
serum alanine
aminotransferase
HDL-C <40 mg/dL
0.84 (0.66, 1.08)
TG >150 mg/dL
1.12 (0.90, 1.39)
tEttinaer et al. (2014)
Kumasi (Ghana), Cape
Town (South Africa),
Victoria (Seychelles),
Kingston (Jamaica),
Maywood, IL (United
States)
Ghana, South Africa,
Seychelles, Jamaica,
United States
2010-2014
Cross-sectional
Modeling the
Epidemiologic Transition
Study (METS)
n = 150
Adults of African descent
from 5 countries of
varying social and
economic development
Blood
Pb measured in venous
whole blood using ICP-
MS
Age at measurement
Mean (SD):
Males: 34.7 (6.0) yr;
Females: 35.2 (6.2) yr
Geometric Mean:
1.55 |jg/dL
Median: 1.66 |jg/dL
75th: 2.6 pg/dL
Max: 31.82 pg/dL
HDL and LDL
cholesterol, blood
pressure,
triglycerides.
Height and weight
were measured by
physical examination.
Fasting glucose was
measured in blood.
Further outcome
assessment details
not provided.
Age at outcome is the
same as age at
exposure assessment
Age, sex, site
location, marital
status, education,
paid employment,
alcohol use, fish
intake
ORs (>1.66 [jg/dL vs.
<1.66 [jg/dL blood Pb)
LDL-C (>2.59 mmol/L)
0.680 (0.289, 1.597)
Triglycerides (>1.7 mmol/L)
0.09 (0.030, 0.250)
HDC-C (<1.03 [males]; <1.29
[females] mmol/L)
1.93 (0.740, 5.020)
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tLiu et al. (2020)
Mexico City
Mexico
Pregnant women
recruited between
1997-1999 and 2001-
2003, follow-up among
offspring began in 2015
Cohort
Early Life Exposure in
Mexico to Environmental
Toxicants (ELEMENT)
n = 369
Mother/child pairs from a
birth cohort study of
pregnant women from 2
public hospitals serving
low to moderate-income
populations
Blood
Maternal Blood Pb
measured in venous
whole blood using
GFAAS
Age at measurement:
Maternal age (SD):
26.7 (5.6) yr
Mean of prenatal blood:
4.3 |jg/dL
Serum lipids
Total cholesterol,
triglycerides, HDL-C,
LDL-C
Age at outcome
Child's age (SD):
13.7 (1.9) yr
Child age, sex, BMI, Change in Z-score (>5 pg/dL
number of siblings at vs. <5 pg/dL blood Pb)
birth, maternal age,
marital status __ . .
education, smoking Triglycerides
history 0.58 (-0.05, 1.20)
Total cholesterol
-0.76 (-1.38, -0.13)
HDL-C
-0.64 (-1.28, 0.01)
LDL-C
-0.96 (-1.59, -0.33)
tKupsco et al. (2019)
Mexico City
Mexico
Maternal blood tested
for metals in 2nd
trimester, children
assessed at age 4-6
Cohort
Research in Obesity,
Growth, Environment and
Social Stressors
(PROGRESS) birth study
n = 548
Mother/child pairs from a
birth cohort study
Blood
Maternal blood Pb
measured second
trimester in venous
whole blood samples
using ICP-MS
Age at measurement
Mean (SD):
28 (5.6) yr
Mean (SD):
3.7 (2.7) |jg/dL
Max: 18 pg/dL
Serum lipids
Triglycerides and non-
HDL cholesterol
Age at outcome:
Mean: 4.8 yr; Range:
4-6 yr
Birth weight,
gestational age, pre-
pregnancy BMI,
education, SES,
parity, environmental
tobacco smoke
Change in Z-score
Triglycerides
0.018 (-0.028, 0.064)
non-HDL-C
-0.015 (-0.058, 0.028)
9-80
-------
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tXu etal. (2017)
United States
1999-2012
Cross-sectional
NHANES
n = 11662
General population; 12-
19 yr old
Blood
Serum lipids
Pb measured in venous Total cholesterol,
whole blood using ICP- triglycerides, HDL-C,
MS
Age at measurement:
12-19 yr
Mean (SD):
1.17 (1.20) |jg/dL
LDL-C
Age at outcome:
12-19 yr
% Increase
Age, gender,
ethnicity, PIR, waist
circumference, serum Tota/ cholesterol
cotinine, and physical
activity
0.6% (-0.1%, 1.3%)
HDL-C
0.3% (-0.5%, 1.1%)
LDL-C
2.3% (0.3%, 4.2%)
Triglycerides
-1.1% (-2.4%, 0.2%)
ALP = alkaline phosphatase; ALT = alanine aminotransferase; AST = aspartate aminotransferase; BMI = body mass index; Cd = cadmium; CI = confidence interval;
CK18 = cytokeratin 18; ELEMENT = Early Life Exposure in Mexico to Environmental Toxicants; FIB-4 = fibrosis-4; GFAAS = graphite furnace atomic absorption spectrometry;
GGT = gamma-glutamyl transferase; Hb = hemoglobin; HDL = high-density lipoprotein; HDL-C = high-density lipoprotein cholesterol; HF = hepatic fibrosis; HS = hepatic
steatosis; ICP-MS = inductively coupled plasma mass spectrometry; KNHANES = Korea National Health and Nutrition Examination Survey; LDL = low-density lipoprotein; LDL-
C = low-density lipoprotein cholesterol; MetS = metabolic syndrome; METS = Modeling the Epidemiologic Transition Study; NAFLD = nonalcoholic fatty liver disease;
NHANES = National Health and Nutrition Examination Survey; NR = not reported; OR = odds ratio; Pb = lead; PIR = poverty-income ratio; PROGRESS = Programming Research
in Obesity, Growth, Environment and Social Stressors; RBC = red blood cell; RR = relative risk; SD = standard deviation; SES = socioeconomic status; SPECT = single photon
emission computed tomography; Q = quartile; TG = thyroglobulin; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
fStudies published since the 2013 Integrated Science Assessment for Pb.
9-81
-------
Table 9-5
Animal toxicological studies of exposure to Pb and hepatic effects
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure Details BLL as Reported (pg/dL)
Endpoints Examined
Berrahal etal. (20111
Rat (Wistar)
0 mg/L Pb Acetate, M,
n = 12-16
50 mg/L Pb Acetate, M,
n = 12-16
PND40, 65
Oral, drinking water
1.76 ± 0.33 |jg/100 mL for
0 mg/L Pb Acetate,
12.67 ± 1.68 |jg/100 mL for
50 mg/mL Pb Acetate - PND 40
2.06 ± 0.35 |jg/100 mL for
0 mg/L Pb Acetate,
7.49 ± 0.78 [jg/100 mL for
50 mg/mL Pb Acetate - PND 65
Plasma Alanine
Aminotransferase (ALT),
Plasma Aspartate
Aminotransferase (AST),
Plasma Alkaline Phosphatase
(ALP)
Li etal. (2017)
Mouse (BALBc)
0 mg/kg Pb Acetate, F,
n = 8
100 mg/kg Pb Acetate, F,
n = 8
Day 29 from Oral, gavage 0.43 ± 0.05 |jg/L for 0 mg/kg Pb
exposure start Acetate
302.20 ± 25.32 pg/L for
100 mg/kg Pb Acetate
Malondialdehyde (MDA)
Levels, Glutathione (GSH),
Glutathione Peroxidase
(GSH-PX), Total Superoxide
Dismutase (T-SOD)
Liu etal. (2013)
Rat (Wistar)
0 ppm Pb, M, n = 10
500 ppm Pb, M, n = 10
Exposure d 75 Oral, drinking water 0.0448 pg/dL for 0 ppm
0.450 pg/dL for 500 ppm
Plasma Alanine
Aminotransferase (ALT),
Plasma Aspartate
Aminotransferase (AST),
GRP78 Protein Levels,
Reactive Oxygen Species
Activity, TBARS Levels, Total
Antioxidant Capacity, ATF6
Protein Levels, ATF4 Protein
Levels, P-IRE1 Protein
Levels, T-IRE1 Protein
Levels, XBP-1 Protein Levels,
P-JNK Protein Levels, JNK
Protein Levels, PI3K Protein
Levels, P-Akt Protein Levels,
T-Akt Protein Levels
9-82
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure Details BLL as Reported (pg/dL)
Endpoints Examined
Long etal. (2016)
Mouse (Kunming)
0% Pb Acetate, M, n = 7
0.2% Pb Acetate, M,
n =21
Six weeks
exposure
Oral, drinking water
36.42 ± 17.48 |jg/L for 0% Pb
Acetate, 214.64 ± 36.24 pg/L for
0.2% Pb Acetate
Plasma Alkaline Phosphatase
(ALP), Plasma Alanine
Aminotransferase (ALT),
Plasma Aspartate
Aminotransferase (AST),
Malondialdehyde (MDA)
Levels, Glutathione (GSH),
Glutathione Peroxidase
(GSH-PX), Total Superoxide
Dismutase (T-SOD),
Apoptosis, Bcl-2 Gene
Expression, Bax Gene
Expression, Bcl-2 Protein
Levels, Bax Protein Levels,
Nrf2 Protein Levels, HO-1
Protein Levels, Gamma-GCS
Protein Levels, Nrf-2 Gene
Expression, HO-1 Gene
Expression, Gamma-GCS
Gene Expression, GRP78
Protein Levels, Grp78 Gene
Expression, Chop Gene
Expression
Andielkovic et al. (2019) Rat (Wistar)
0 mg Pb Acetate per kg
bw, M, n = 8
150 mg Pb Acetate per
kg bw, M, n = 6
Dehydrogenase (LDH),
Malondialdehyde (MDA)
Levels, Advanced Oxidation
Protein Products Level
(AOPP), Total Thiol (-SH)
Groups Level, Prooxidative-
Antioxidative Balance (PAB),
Total Superoxide Dismutase
(T-SOD)
24 h posttreatment Oral, gavage
25 pg/L for 0 mg Pb Acetate per
kg bw, 290 pg/L for 150 mg Pb
Acetate per kg bw
Plasma Aspartate
Aminotransferase (AST),
Plasma Alanine
Aminotransferase (ALT),
Plasma Alkaline Phosphatase
9-83
-------
Study
Species (Stock/Strain), Timing^of Exposure Details BLL as Reported (Mg/dL) Endpoints Examined
Dumkova et al. (2017)
Mouse (ICR) Week 6 of
0 particles/cm3, F, n = 10 exPosure
1.23 x 10s particles/cm3,
F, n = 10
Inhalation
1.1 |jg/dL for 0 particles/cm3,
13.2 |jg/dL for 1.23 * 10®
particles/cm3, F, n = 10
Histopathology, Proliferating
Cell Nuclear Antigen (PCNA)
Immunohistochemistry,
Apoptotic Cells (TUNEL-
Positive), Na-KATPase
Expression
Barkaoui et al. (2020)
Rat (Wistar)
0 g/L Pb Acetate, M,
n =6
1 g/L Pb Acetate, M,
n =6
Exposure day 30
Oral, drinking water 11.1 ±0.12 pg/dL for 0 g/L Pb
Acetate
23.8 ± 0.912 |jg/dL fori g/L Pb
Acetate
GSH, CAT, T-SOD, GSH-PX,
MDA Levels, Histopathology,
CAT qRT-PCR, GPx qRT-
PCR, SOD qRT-PCR, NF-kB
qRT-PCR, IL-6 qRT-PCR,
TNF-alpha qRT-PCR
Gao et al. (2020)
Rat (Sprague Dawley)
0 mg/kg bw, Pb2+, M/F,
n = 10
5 mg/kg bw, Pb2+, M/F,
n = 10
Four weeks Oral, gavage 0.02 mg/kg for 0 mg/kg bw,
postexposure Pb2+,
0.1 ± 0.03 mg/kg for 5 mg/kg
bw, Pb2+
T-SOD, CAT, MDA Levels,
GSH, Histopathology, Plasma
AST, Plasma ALT, Cr, BUN
Dumkova et al. (2020b)
Mouse (Not Specified)
0 |jg/m3 PbO NPs, F,
n = NR, 2, 6, 11 wk
78.0 |jg/m3 PbO NPs, F,
n = NR, 6 wk followed by
0 |jg/m3 PbO NPs, 5 wk
78.0 |jg/m3 PbO NPs, F,
n = NR, 2, 6, 11 wk
Exposure week 2,
6, 11
Inhalation
0 |jg/dL for 0 |jg/m3 PbO NPs,
F, n = NR, 2, 6, 11 wk
2.7 |jg/dL for 78.0 |jg/m3 PbO
NPs, F, n = NR, 6 wk followed
by 0 |jg/m3 PbO NPs, 5 wk
10.4 ug/dL for 78.0 ug/m3 PbO
NPs-2 wk
14.8 ug/dL for 78.0 ug/m3 PbO
NPs-6 wk
17.4 |jg/dL for 78.0 |jg/m3 PbO
NPs-11 wk
Plasma Alkaline Phosphatase
(ALP), Plasma Alanine
Aminotransferase (ALT),
Plasma Aspartate
Aminotransferase (AST), Cr
9-84
-------
Study
Species (Stock/Strain),
n, Sex
Timing of
Exposure
Exposure Details BLL as Reported (pg/dL)
Endpoints Examined
Dumkova et al. (2020a)
Mouse (CD1), (ICR)
0 |jg/m3 Pb(N03)2 NPs,
F, n = 10-3d, 2, 6,
11 wk
68.6 |jg/m3 Pb(N03)2
NPs, F, n = 10 - 3 d, 2, 6
11 wk
68.6 |jg/m3 Pb(N03)2
NPs, F, n = 10 - 6 wk,
followed by 0 |jg/m3
Pb(N03)2 NPs - 5 wk
Exposed 3 d,
11 wk
2, 6,
Inhalation
0 pg/dL for 0 |jg/m3 - all groups
3.1 pg/dL for 68.6 |jg/m3 -3d
4.0 pg/dL for 68.6 |jg/m3 - 2 wk
4.7 pg/dL for 68.6 |jg/m3 - 6 wk
8.5 pg/dL for 68.6 |jg/m3 -11 wk
1.0 pg/dL for 68.6 |jg/m3 - 6 wk
followed by 0 |jg/m3 - 5 wk
Histopathology, NF-kB qRT-
PCR, TNF-alpha qRT-PCR,
IL-1 alpha, IL-1 beta, IL-6
qRT-PCR, TGFbetal, Plasma
Alkaline Phosphatase (ALP)
Laamech et al. (2017)
Mouse (IOPS)
0 mg/kg body weight/day
Pb Acetate, M, n = 12
5 mg/kg body weight/day
Pb Acetate, M, n = 12
Exposure day 40 Oral, gavage
0.010 pg/mLforO mg/kg body
weight/day Pb Acetate,
0.18 |jg/ml_ for 5 mg/kg body
weight/day Pb Acetate
Histopathology, Plasma
Alanine Aminotransferase
(ALT), Plasma Aspartate
Aminotransferase (AST),
Total Cholesterol (TC), Total
Bilirubin (TB),
Malondialdehyde (MDA)
Levels, Protein Carbonyl
(PCO), Glutathione (GSH),
Catalase, Total Superoxide
Dismutase (T-SOD),
Glutathione Peroxidase
(GSH-PX)
ALP = alkaline phosphatase; ALT = alanine aminotransferase; AOPP = advanced oxidation protein products; AST = aspartate aminotransferase; BUN = blood urea nitrogen;
BLL = blood lead levels; CAT = catalase; Cr = chromium; D = day(s); GSH = glutathione; GSH-PX = glutathione peroxidase; LDH = lactate dehydrogenase; h = hour;
MDA = malondialdehyde; NF-kB = nuclear factor kappa B; NP = nanoparticle; PAB = prooxidative-antioxidative balance; Pb = lead; PCNA = proliferating cell nuclear antigen;
PCO = protein carbonyl; PND = postnatal day; qRT-PCR = real-time quantitative reverse transcription-polymerase chain reaction; TB = total bilirubin; TBARS = thiobarbituric acid
reactive substance; TC = total cholesterol; T-SOD = total superoxide dismutase; wk = week(s).
9-85
-------
Table 9-6 Epidemiologic studies of exposure to Pb and metabolic effects
sffiySS? PopBlon Exposure Assessment Outcome Contenders E«ec. Estimates and 95%
Diabetes and Insulin Resistance - Adults
tMoon (2013)
Korea
2007-2012
Cross-Sectional
KNHANES
n = 3,184
Adults aged
>30 yr
Blood
Pb was measured in venous
whole blood using GFAAS
Age at measurement
Mean (SD):
No diabetes: 49.4 (12.4) yr
Diabetes: 58.8 (10.9) yr
Geometric Mean (SD):
No diabetes:
2.41 (1.52) |jg/dL
Diabetes:
2.47 (1.59) |jg/dL
Diabetes, HOMA-IR, HOMA-I3 Age, sex, region,
(%), fasting insulin (mlU/L)
Age at outcome is the same
as age at exposure
assessment
smoking, alcohol
consumption, regular
exercise, BMI (sex-
stratified analyses
only)
OR (95% CI) for prevalent
diabetes across blood Pb
quartiles:
Q1 (GM 1.43 pg/dL):
Reference
Q2 (GM 2.13 pg/dL):
0.91 (0.64, 1.29)
Q3 (GM 2.74 pg/dL):
0.76 (0.53, 1.09)
Q4 (GM 4.08 pg/dL):
0.75 (0.52, 1.08);
Change in HOMA-IR,
HOMA-IJ, and Fasting
Insulin per unit increase
in log-blood Pb
log(HOMA-IR)
Men: -0.04 (-0.10, -0.02),
Women: -0.04 (-0.09,
-0.01)
log(HOMA-IJ)
Men: -0.05 (-0.11, 0.01),
Women: -0.05 (-0.10,
0.01)
Fasting insulin (mlU/L)
Men: -0.53 (-1.23, 0.16)
Women: -0.27 (-1.00,
0.46)
9-86
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tHansen et al. (2017)
Nord-Trondelag
Health Study
Nord-Trondelag
(HUNT3)
County
n = 883
Norway
2006-2008
Adults aged
Nested Case-Control
>20 yr. Cases
(n = 128) were
HUNT3
participants
diagnosed with
diabetes.
Controls
(n = 755) were
age- and sex-
matched HUNT3
participants
without diabetes.
Blood
Pb was measured in venous
whole blood using ICP-MS
Age at measurement
Mean (SD):
Cases: 61.4 (14.1) yr
Controls: 65.2 (10.3) yr
Median (10th—90th
percentile):
Cases: 19.9 (10.8-38.0) pg/L
Controls: 19.4 (11.0-37.2)
pg/L
Type 2 diabetes
Individuals were screened for
diabetes at a physical
examination using an oral
glucose tolerance test.
Diagnosis with type 2
diabetes was defined as
having fasting serum glucose
>7.0 mmol/L and/or 2 h
glucose >11.1 mmol/L as well
as glutamic acid
decarboxylase antibodies
(GADA) <0.08 ai.
Age at outcome is the same
as age at exposure
assessment
Age, sex, BMI, waist-
to-hip ratio,
education, income,
smoking, family
history of diabetes
OR (95% CI) for prevalent
type 2 diabetes Q4 vs.
Q1: 1.12 (0.58, 2.16)
tSimic etal. (2017)
Norway
2006-2008
Nested Case-Control
Nord-Trondelag
Health Study
(HUNT3)
n = 945
Adults aged
>20 yr. Cases
(n = 270) were
HUNT3
participants
diagnosed with
type 2 diabetes.
Controls
(n = 615) were
age- and sex-
matched
participants
without diabetes.
Blood
Pb was measured in venous
whole blood using ICP-MS
Age at measurement
Mean (SD):
Cases: 59.2 (12.2) yr
Controls: 65.4 (10.6) yr
Median (10th-90th
percentile):
Cases: 16.4 (9.7-35.2) pg/L
Controls: 20.2 (11.2-37.9)
pg/L
Type 2 diabetes
Type 2 diabetes was defined
as self-reported diabetes
excluding type I diabetes as
indicated by GADA index,
measured in blood at a
physical examination.
Age at outcome is the same
as age at exposure
assessment
BMI, waist-to-hip
ratio, first-degree
family history of
diabetes, smoking
habits, area,
education, economic
status, alcohol
consumption, blood
calcium
OR (95% CI) for prevalent
type 2 diabetes Q4 vs.
Q1: 0.24 (0.13, 0.47)
9-87
-------
sSEE.S? Popul^on Exposure Assessment Outcome Confounders Es.imetes end 95%
Diabetes and Insulin Resistance - Adolescents
tLiu et al. (2020)
Mexico City
Maternal enrollment:
1997-1999 and
2001-2003
Child follow-up: 2015
Prospective Birth
Cohort
Early Life
Exposure in
Mexico to
Environmental
Toxicants
(ELEMENT)
Study
n = 369
Adolescents
aged 10-18 yr
Blood
Maternal Pb (1st trimester)
was measured in venous
whole blood using GFAAS
Age at measurement
Mean maternal age in 1st
trimester of pregnancy (SD):
26.7 (5.6) yr
Geometric Mean (95% CI):
4.3 (4.0, 4.6) |jg/dL
Fasting serum glucose Z-
score (mg/dL), HOMA-IR Z-
score
Serum fasting glucose
(mg/dL) was measured using
an enzymatic method. Serum
insulin (|jU/mL) was
measured using
immunoturbidimetric assay.
HOMA-IR was calculated as
insulin (|jU/mL)*glucose
(mg/dL)/405.
Age at outcome
Mean child age (SD):
13.7 (1.9) yr
Child age, sex, BMI z-
score, number of
siblings at birth,
maternal age, marital
status, education,
smoking history
Change in mean fasting
glucose and HOMA-IR Z-
scores for maternal blood
Pb >5 |jg/dL vs. maternal
blood Pb <5 [jg/dL
Fasting glucose z-score
All: -0.05 (-0.69, 0.60)
Boys: -0.05 (-0.34, 0.25)
Girls: -0.06 (-0.35, 0.23)
HOMA-IR z-score
All: -0.11 (-0.63, 0.42)
Boys: -0.04 (-0.28, 0.20)
Girls: 0.04 (-0.19, 0.27)
9-88
-------
sSEE.S? Popul^on Exposure Assessment Outcome Confounders Es.imetes end 95%
Metabolic Syndrome (MetS) and Its Components
tMoon (2014)
Korea
2007-2012
Cross-Sectional
KNHANES
n = 3,950
Adults aged
>20 yr
Blood
Pb measured in venous
whole blood using GFAAS.
Age at measurement
Mean (SD):
No MetS: 42.7 (14.6) yr;
MetS: 54.4 (12.8) yr
Mean (SD)
No MetS: 2.08 (1.00) pg/dL;
MetS: 2.50 (1.01) pg/dL
Metabolic syndrome
MetS was defined as meeting
at least 3 of the following: (1)
elevated blood pressure (SBP
>130 mmHg or DBP
>85 mmHg or current use of
blood pressure medication),
(2) low HDL cholesterol
(<40 mg/dL in women or
<50 mg/dL in men), (3)
elevated serum triglycerides
(>150 mmHg) or current use
of antidyslipidemia
medication, (4) elevated
fasting plasma glucose levels,
(5) abdominal obesity (waist
circumference >90 cm in men
or >85 cm in women).
Age, sex, region,
smoking, alcohol
consumption, regular
exercise, BMI
OR (95% CI) for prevalent
MetS across blood Pb
quartiles
Q1 (GM 1.23 pg/dL):
Reference
Q2 (GM 1.90 pg/dL):
0.84 (0.62, 1.13)
Q3 (GM 2.51 pg/dL):
1.21 (0.90, 1.62)
Q4 (GM 3.79 pg/dL):
1.07 (0.79, 1.45)
Age at outcome is the same
as age at exposure
assessment
9-89
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tRhee et al. (2013)
Korea
2008
Cross-Sectional
KNHANES
n = 1,405
Nationally
representative
survey of Korean
adults
Blood
Pb was measured in venous
whole blood using GFAAS
Age at measurement
Mean (SD):
No MetS: 40.3 (13.7) yr
MetS: 47.1 (13.3) yr
Median (25th—75th):
2.35 (1.74-3.06) pg/dL
75th: 3.06 pg/dL
Max: 19.43 pg/dL
MetS, abdominal
circumference, triglycerides,
HDL cholesterol, fasting
glucose
MetS was defined using the
Modified National Cholesterol
Education Program Adult
Treatment Panel III Criteria,
with the exception of waist
circumference measurement
cut-offs of >90 cm for males
and >85 cm for females
based on criteria from the
Korean Society for the Study
of Obesity. TC, triglycerides,
HDL cholesterol, and fasting
plasma glucose were
assessed using an automated
analyzer with enzymatic
assays. Abdominal
circumference was measured
by a professional.
Age at outcome is the same
as age at exposure
assessment
Age, sex, smoking,
education, TC,
creatinine, AST, ALT,
fasting serum insulin
OR for MetS prevalence
across log-transformed
Pb quartiles
Q1 (<1.73 pg/dL):
Reference
Q2 (1.74-2.35 pg/dL):
1.56 (0.90, 2.71)
Q3 (2.35-3.06 pg/dL):
1.63 (0.94, 2.83)
Q4 (>3.07 pg/dL):
2.57 (1.46, 4.51)
Change in outcomes per
unit increase in log-
transformed Pb
Abdominal circumference
0.051 (-0.001, 0.107) cm
Triglycerides
0.080 (0.023, 0.137) mg/dL
HDL Cholesterol
0.033 (-0.020, 0.086)
mg/dL
Fasting Glucose
0.019 (-0.029, 0.067)
mg/dL
9-90
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tBulka etal. (2019)
United States
2011-2014
Cross-Sectional
NHANES
n = 1,088
Nationally
representative
survey of U.S.
adults
Blood
Pb was measured in venous
whole blood using ICP-MS
Age at measurement:
20-60 yr
Mean (SD)
NHANES 2011-2012:
1.17 (0.04) |jg/dL;
NHANES 2013-2014:
1.00 (0.03) |jg/dL
MetS, triglycerides, HDL
cholesterol, blood glucose,
abdominal obesity
MetS was defined as meeting
at least 3 of the following: (1)
elevated blood pressure (SBP
>130 mmHg or DBP
>85 mmHg or current use of
blood pressure medication),
(2) low HDL cholesterol
(<40 mg/dL in women or
<50 mg/dL in men), (3)
elevated serum triglycerides
(>150 mmHg) or current use
of antidyslipidemia
medication, (4) elevated
fasting plasma glucose levels,
(5) abdominal obesity (waist
circumference >90 cm in men
or >85 cm in women). Waist
circumference (cm) was
measured at the physical
examination by a trained
professional. Serum HDL
(|jg/dL), triglycerides (mg/dL),
and blood glucose (mg/dL)
were measured in blood
samples obtained in the
morning following an
overnight fast.
Age at outcome:
20-60 yr
Age, race/ethnicity,
family poverty-income
ratio, total caloric
intake, educational
attainment, smoking
status, average
number of drinks per
day past year,
physical activity
status, survey cycle,
BMI (excluding
abdominal obesity
analysis), serum
cotinine
PRs for outcomes across
Pb quartiles
MetS
Q1 (0.18-0.70 pg/dL):
Reference
Q2 (0.71-1.05 pg/dL):
0.90 (0.73, 1.11)
Q3 (1.06-1.63 pg/dL):
0.84 (0.69, 1.05)
Q4 (1.64-15.98 pg/dL):
0.81 (0.64, 1.03)
High Triglycerides
Q1
Q2
Q3
Q4
Reference
0.85 (0.72, 0.99)
0.76 (0.64, 0.92)
0.82 (0.67, 1.01)
Low HDL
Q1: Reference
Q2
Q3
Q4
0.90 (0.76, 1.07)
0.79 (0.65, 0.97)
0.73 (0.59, 0.89)
High Glucose
Q1: Reference
Q2
Q3
Q4
1.03 (0.86, 1.23)
0.86 (0.68, 1.08)
0.95 (0.77, 1.17)
Abdominal Obesity
Q1: Reference
Q2
Q3
Q4
0.93 (0.82, 1.07)
0.91 (0.80, 1.04)
0.66 (0.56, 0.78)
9-91
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tShimetal. (2019)
Korea
2012-2014
Cross-Sectional
Korean National
Environmental
Health Survey II
(KNEHS)
n = 5,251
Nationally
representative
survey of adults
in Korea
Blood
Pb was measured in venous
whole blood using GFAAS
Age at measurement
Mean (SE):
No MetS: 49.87 (0.22) yr
MetS: 61.59 (0.50) yr
Geometric Mean (SE)
No MetS: 0.71 (0.48) pg/dL
MetS: 0.76 (0.49) pg/dL
MetS
MetS was defined as meeting
at least 3 of the following: (1)
elevated blood pressure (SBP
>130 mmHg or DBP
>85 mmHg or current use of
blood pressure medication),
(2) low HDL cholesterol
(<40 mg/dL in women or
<50 mg/dL in men), (3)
elevated serum triglycerides
(>150 mmHg) or current use
of antidyslipidemia
medication, (4) elevated
fasting plasma glucose levels,
(5) abdominal obesity (waist
circumference >90 cm in men
or >85 cm in women).
Age, sex, education,
income, marital
status, aspartate
aminotransferase,
alanine
aminotransferase
ORs for MetS prevalence
across blood Pb quartiles
Q1
Q2
Q3
Q4
Reference
0.94 (0.72, 1.24)
1.00 (0.76, 1.31)
0.86 (0.65, 1.14)
Quartile levels NR
Age at outcome is the same
as age at exposure
assessment
tWen et al. (2020) N = 2444
Taiwan
June 2016-
September2018
Cross-Sectional
General
population
Blood
Pb was measured in venous
whole blood using ICP-MS
Age at measurement:
Mean (SD): 55.1 (13.2) yr
MetS
Age, sex, TC, LDL
cholesterol,
hemoglobin, eGFR,
uric acid
OR MetS prevalence per
log unit increase in blood
Pb:
0.86 (0.61, 1.20)
Mean: 1.6 pg/dL
9-92
-------
sSEE.S? Popul^on Exposure Assessment Outcome Confounders Estimates ,„d 95%
MetS was defined as meeting
at least 3 of the following: (1)
elevated blood pressure (SBP
>130 mmHg or DBP
>85 mmHg or current use of
blood pressure medication),
(2) low HDL cholesterol
(<40 mg/dL in women or
<50 mg/dL in men), (3)
elevated serum triglycerides
(>150 mmHg) or current use
of antidyslipidemia
medication, (4) elevated
fasting plasma glucose levels,
(5) abdominal obesity (waist
circumference >90 cm in men
or >85 cm in women).
Age at outcome:
Mean (SD): 55.1 (13.2) yr
9-93
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tLee and Kim (2016) KNHANES
n = 9,880
Korea
2007-2012 Korean adults
Cross-Sectional aged >20 yr
Blood
Pb measured in venous
whole blood using GFAAS
Age at measurement
Mean (SD):
Males
No MetS: 43.5 (0.23) yr
MetS: 48.7 (0.48) yr
Females
No MetS: 43.5 (0.25) yr
MetS: 51.4 (0.60) yr
Geometric Mean (SD):
Males
No MetS: 2.57 (0.02) pg/dL
MetS: 2.64 (0.04) pg/dL
Females
No MetS: 1.86 (0.01)
MetS: 1.92 (0.04) pg/dL
MetS, waist circumference
(cm), serum HDL (mg/dL),
serum triglycerides (mg/dL),
blood glucose (mg/dL)
MetS was defined as meeting
at least 3 of the following: (1)
elevated blood pressure (SBP
>130 mmHg or DBP
>85 mmHg or current use of
blood pressure medication),
(2) low HDL cholesterol
(<40 mg/dL in women or
<50 mg/dL in men), (3)
elevated serum triglycerides
(>150 mmHg) or current use
of antidyslipidemia
medication, (4) elevated
fasting plasma glucose levels,
(5) abdominal obesity (waist
circumference >90 cm in men
or >85 cm in women). Waist
circumference (cm) was
measured at the physical
examination by a trained
professional. Serum HDL
(pg/dL), triglycerides (mg/dL),
and blood glucose (mg/dL)
were measured in blood
samples obtained in the
morning following an
overnight fast.
Age at outcome is the same
as age at exposure
assessment
Age, BMI, residence
area, education level,
smoking and drinking
status, exercise, AST,
ALT
OR (95% CI) for outcomes
across blood Pb tertiles
MetS prevalence
T1 (<2.20 pg/dL):
Reference
T2 (2.20-3.01 pg/dL):
1.032 (0.788, 1.352)
T3 (>3.01 pg/dL):
0.817 (0.626, 1.065)
Waist circumference
(>85 cm)
T1
T2
T3
Reference
1.11 (0.83, 1.50)
1.11 (0.81, 1.51)
Serum HDL (<40 mg/dL)
T1
T2
T3
Reference
1.00 (0.80, 1.24)
0.76 (0.59, 0.97)
Serum triglycerides
(>150 mg/dL)
T1: Reference
T2: 1.13 (0.93, 1.39)
T3: 1.08 (0.87, 1.33)
Blood glucose
(>100 mg/dL)
T1: Reference
T2: 0.83 (0.68, 1.02)
T3: 1.04 (0.85, 1.28)
9-94
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tLee and Kim (2013) KNHANES
n = 7,559
Korea
2005-2010 Korean adults
Cross-Sectional aged >20 yr
Blood
Pb measured in venous
whole blood using GFAAS
Age at measurement
Mean (SD):
No MetS: 42.3 (0.29) yr
MetS: 48.4 (0.57) yr
Geometric Mean (SD):
No MetS: 2.734 (0.024) pg/dL
MetS: 2.957 (0.049) pg/dL
MetS, waist circumference,
serum HDL, serum
triglycerides, blood glucose
MetS was defined as meeting
at least 3 of the following: (1)
elevated blood pressure (SBP
>130 mmHg or DBP
>85 mmHg or current use of
blood pressure medication),
(2) low HDL cholesterol
(<40 mg/dL in women or
<50 mg/dL in men), (3)
elevated serum triglycerides
(>150 mmHg) or current use
of antidyslipidemia
medication, (4) elevated
fasting plasma glucose levels,
(5) abdominal obesity (waist
circumference >90 cm in men
or >85 cm in women). Waist
circumference (cm) was
measured at the physical
examination by a trained
professional. Serum HDL
(pg/dL), triglycerides (mg/dL),
and blood glucose (mg/dL)
were measured in blood
samples obtained in the
morning following an
overnight fast.
Age at outcome is the same
as age at exposure
assessment
Age, BMI, residence
area, education level,
smoking and drinking
status, exercise,
serum aspartate
aminotransferase,
serum alanine
aminotransferase
OR (95% CI) for outcomes
across blood Pb tertiles
MetS Prevalence
T1 (<2.362 pg/dL):
Reference
T2 (>2.362-3.282 pg/dL):
1.267 (0.950, 1.690)
T3 (>3.282 pg/dL:
0.984 (0.735, 1.317)
Waist circumference
(>85 cm)
T1
T2
T3
Reference
1.04 (0.75, 1.45)
0.89 (0.64, 1.24)
Serum HDL (<40 mg/dL)
T1: Reference
T2
T3
0.98 (0.79, 1.23)
0.96 (0.77, 1.20)
Serum triglycerides
(>150 mg/dL)
T1: Reference
T2
T3
1.01 (0.82, 1.24)
1.07 (0.87, 1.32)
Blood glucose
(>100 mg/dL)
T1: Reference
T2: 1.00 (0.81, 1.24)
T3: 1.14 (0.91, 1.44)
9-95
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tWanqetal. (2018c) NHANES
n = 9537
United States
2003-2014
Cross-Sectional
NHANES
participants aged
20+,2003-2014
Blood
Pb was measured in venous
whole blood using ICP-DRC-
MS
Age at measurement
Mean (SD): 49.2 (18.0) yr
Geometric mean (SD):
1.32 (2.00) |jg/dL
Waist circumference (cm)
Waist circumference (cm)
was measured during minimal
respiration to the nearest
0.1 cm at the level of the iliac
crest at the time of NHANES
physical examination.
Age at outcome:
Mean (SD): 49.2 (18.0) yr
Age, sex,
race/ethnicity,
education, smoking
status, physical
activity, NHANES
cycle, and urinary
creatinine
Change in waist
circumference (cm) per 1-
SD increase in log(10)-
transformed Pb (SD NR):
0.008 (-0.010, -0.006)
tPeters et al. (2012) Normative Aging Blood, Bone
Serum lipids
United States
Blood Pb measured
between 1999-2008;
Serum lipids
measured 3 to 4 yr
after blood Pb
Cohort
Study
n = 426
Older male
Veterans
Blood Pb measured in venous Triglycerides, HDL-C
whole blood using GFAAS
Age at outcome:
Mean: 4.01 ± 2.30 |jg/dL 3 to 4 yr after blood Pb
Age at baseline, yr
between baseline and
outcome, education,
BMI, alcohol intake,
smoking status, pack-
yr of smoking,
hypertension status,
and statin use
ORs
Low HDL-C (<40 mg/dL):
0.899 (0.804, 1.004)
High Triglycerides
(>200 mg/dL):
0.993 (0.874, 1.129)
tEttinaer et al. (2014)
Kumasi, Ghana; Cape
Town, South Africa;
Victoria, Seychelles;
Kingston, Jamaica;
Maywood, Illinois
(United States)
2010-2014
Prospective Cohort
Modeling the
Epidemiologic
Transition Study
(METS)
n = 150
Adults of African
descent from 5
countries of
varying social
and economic
development
Blood
Pb was measured in venous
whole blood using DRC-ICP-
MS
Age at measurement
Mean (SD):
Males: 34.7 (6.0) yr
Females: 35.2 (6.2) yr
Geometric Mean (95% CI):
1.55 (1.30, 1.85) |jg/dL
Waist Circumference >94 cm
(males) or >80 cm (females),
Fasting Glucose >100 mg/dL
Fasting glucose was
measured in blood. Further
outcome assessment details
not provided.
Age at outcome is the same
as age at exposure
assessment
Age, sex, site
location, marital
status, education,
paid employment,
alcohol use, fish
intake, percent body
fat
ORs for blood Pb above
the median (1.66 (jg/dL)
vs. below the median
Waist Circumference
[ >94 cm (m) or >80 cm (f)]
4.53 (1.06, 19.48)
Fasting Glucose
(>100 mg/dL)
4.99 (1.97, 12.69)
Median (95% CI):
1.66 (1.34, 1.93) |jg/dL
9-96
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
75th: 2.6 pg/dL
Max: 31.82 pg/dL
Body Weight
tWanqetal. (2018a)
China
2014
Cross-Sectional
SPECT-China
n = 3922
Chinese citizens
aged >18 yr who
had lived in their
current area for
6+ mo
Blood
Pb was measured in venous
whole blood using GFAAS
Age at measurement:
Mean (SD):
Normal weight subjects:
50.9 (13.9) yr
Overweight subjects:
54.0 (12.3) yr
Obese subjects:
56.2 (11.2) yr
Median
(25th-75th percentiles)
Normal weight:
3.9 (2.6, 5.6) pg/dL
Overweight subjects:
4.3 (2.9, 6.1) pg/dL
Obese subjects:
4.4 (2.7, 6.2) pg/dL
BMI (kg/m2)
BMI was calculated as weight
(kg) divided by squared
height (m2). Overweight
(including obese) was defined
as BMI >25 kg/m2.
Age at outcome is the same
as age at exposure
assessment
Age, sex, economic
status, rural/urban
residence, current
smoking, diabetes,
hypertension,
dyslipidemia
OR (95% CI) for
overweight or obese (BMI
>25 kg/m2) across blood
Pb quartiles
Q1 (<2.69 pg/dL):
Reference
Q2 (2.69-4.01 pg/dL):
1.09 (0.89, 1.33)
Q3 (4.01-5.60 pg/dL):
1.15 (0.94, 1.40)
Q4 (>5.60 pg/dL):
1.40 (1.14, 1.71)
9-97
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tEttinqer et al. (2014)
Kumasi, Ghana; Cape
Town, South Africa;
Victoria, Seychelles;
Kingston, Jamaica;
Maywood, Illinois
(United States)
2010-2014
Prospective Cohort
Modeling the
Epidemiologic
Transition Study
(METS)
n = 150
Adults of African
descent from 5
countries of
varying social
and economic
development
Blood
Pb was measured in venous
whole blood using DRC-ICP-
MS
Age at measurement
Mean (SD):
Males: 34.7 (6.0) yr
Females: 35.2 (6.2) yr
Geometric Mean (95% CI):
1.55 (1.30, 1.85) |jg/dL
Overweight (BMI >25), Obese Age, sex, site
(BMI >30)
Height and weight were
measured by physical
examination.
Age at outcome is the same
as age at exposure
assessment
location, marital
status, education,
paid employment,
alcohol use, fish
intake, percent body
fat
ORs for blood Pb above
the median (1.66 (jg/dL)
vs. below the median
Overweight (BMI >25)
0.88 (0.31, 2.51)
Obese (BMI >30)
2.70 (0.75, 9.75)
Median (95% CI):
1.66 (1.34, 1.93) |jg/dL
75th: 2.6 pg/dL
Max: 31.82 pg/dL
9-98
-------
Reference and
Study Design
Study
Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
CIs
tGuoetal. (2019) N = 145
China
2015
Cross-Sectional
Blood
BMI (kg/m2)
Adult men Pb was measured using ICP- Age outcome
recruited through MS Mean (SD): 39 (12) yr
a physical
examination Age measurement
center Mean (SD): 39(12) yr
Age
Change in BMI (kg/m2)
per log increase in blood
Pb:
0.05 (-3.64, 3.74)
Mean (SD): 8.5 (3.8) pg/dL;
Median: 7.9 pg/dL
75th: 10.8 pg/dL
Max: 28.2 pg/dL
ALT = alanine aminotransferase; AMT = ;AST = aspartate aminotransferase; BMI = body mass index; CI = confidence interval; DBP = diastolic blood pressure; DRC-ICP-
MS = dynamic reaction cell for inductively coupled plasma mass spectrometry; eGFR = estimated glomerular filtration rate; ELEMENT = Early Life Exposure in Mexico to
Environmental Toxicants; GADA = glutamic acid decarboxylase antibodies; GFAAS = graphite furnace atomic absorption spectrometry; GM = geometric mean; HDL = high-
density lipoprotein; HDL-C = high-density lipoprotein cholesterol; HOMA-IR = Homeostatic Model Assessment for Insulin Resistance; HOMA- (B = HOMA of (B-cell function; ICP-
MS = inductively coupled plasma mass spectrometry; KNHANES = Korea National Health and Nutrition Examination Survey; MetS = metabolic syndrome; METS = Modeling the
Epidemiologic Transition Study; NR = not reported; OR = odds ratio; Pb = lead; SBP = systolic blood pressure; SD = standard deviation; SPECT = single photon emission
computed tomography; TC = total cholesterol; Q = quartile; yr = year.
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
fStudies published since the 2013 Integrated Science Assessment for Pb.
9-99
-------
Table 9-7
Animal toxicological studies of exposure to Pb and metabolic effects
Study
Species (Stock/Strain), Timing of Exposure
n, Sex Exposure Details
BLL as Reported (pg/dL)
Endpoints Examined
Faulk et al. Mouse (Agouti), 0.0 ppm
(2014) Pb, M/F, n = 30
2.1 ppm Pb, M/F, n = 28
16 ppm Pb, M/F, n = 33
32 ppm Pb, M/F, n = 29
(Longitudinal phenotypic
measures were taken
from a total of 120 a/a
mice, on average 2.7
mice per litter)
Mo 3, 6, 9
Oral, drinking
water
Mean maternal BLL,
tested at weaning, were
below the LOD for the
control group, and 4.1
(61.3) |jg/dL, 25.1 (67.3)
|jg/dL, and 32.1 (611.4)
|jg/dL in the three
exposure groups,
2.1 ppm, 16 ppm, and
32 ppm, respectively
Oxygen Consumption, CO2 Production, Food Intake, Body
Weight, Body Fat
Rahman et al.
(2018)
Rat (Wistar)
0% Pb Acetate, M/F,
n = 37
0.2% Pb Acetate, M/F,
n = 38
PND 21, 30 Oral, drinking 2.2 ± 0.07 pg/dL for 0% Serum 25(OH)D, Serum 1,25(OH)2D, Hepatic 25-
water Pb Acetate,
12.4 ± 3.3 pg/dL for 0.2%
Pb Acetate - PND 21
3.3 ± 1.7 pg/dL for 0% Pb
Acetate, 22.7 ± 6.0 pg/dL
for 0.2% Pb Acetate -
PND 30
Hydroxylase Protein Levels, Hepatic 25-Hydroxylase
Immunohistochemistry
9-100
-------
Study
Species (Stock/Strain), Timing of Exposure
n, Sex Exposure Details
BLL as Reported (pg/dL)
Endpoints Examined
Zhou et al. Rat (Sprague Dawley),
(2018) 0%Pb Acetate, M,
n = 20
0.5% Pb Acetate, M,
n = 20
1% Pb Acetate, M,
n = 20
2% Pb Acetate, M,
n = 20
PND 52 Oral, drinking 11.4 |jg/L for 0%
water 147 |jg/L for 0.5%
226 |jg/L for 1 %
289 |jg/L for 2%
Cholesterol Content, mRNA level of SREBP2 in the
cortex, mRNA level of SREBP2 in the hippocampus,
mRNA level of LDL-R in the cortex, mRNA level of HMG-
CR in the hippocampus, mRNA level of HMG-CR in the
cortex, protein level of SREBP2 in the cortex, mRNA level
of LDL-R in the hippocampus, protein level of HMG-CR in
the cortex, protein level of LDL-R in the cortex, protein
level of SREBP2 in the hippocampus, protein level of
HMG-CR in the hippocampus, protein level of LDL-R in
the hippocampus, immunohistochemistry ofSREBP2 in
the cortex, immunohistochemistry of HMG-CR in the
cortex, immunohistochemistry of LDL-R in the cortex,
immunohistochemistry of SREBP2 in the hippocampus,
immunohistochemistry of HMG-CR LDL-R in the
hippocampus, immunohistochemistry of LDL-R in the
hippocampus, mRNA level of LXR-a in the cortex, mRNA
level of ABCA1 in the cortex, mRNA level of LXR-a in the
hippocampus, mRNA level of ABCA1 in the hippocampus,
protein level of LXR-a in the cortex, protein level of
ABCA1 in the cortex, protein level of LXR-a in the
hippocampus, protein level of ABCA1 in the hippocampus
ABCA1 = ATP-binding cassette transporter ABCA1 (member 1 of human transporter sub-family ABCA); BLL = blood lead level; C02 = carbon dioxide; F = female; HMG-CR = 3-
Hydroxy-3-Methylglutaryl-Coenzyme A Reductase; LDL-R = low-density lipoprotein receptor; LOD = limit of detection; LXR-a = liver X receptor alpha; M = male; mo = month;
mRNA = messenger ribonucleic acid; Pb = lead; PND = postnatal day; SREBP2 = Sterol Regulatory Element Binding Transcription Factor 2.
9-101
-------
Table 9-8 Animal toxicological studies of exposure to Pb and gastrointestinal effects
Study
Species (Stock/Strain),
n, Sex
Timing of Exposure
Exposure Details
BLL as Reported
(Hg/dL)
Endpoints Examined
Kosik-Boaacka et al.
(2011)
Rat (Wistar), Control
(distilled water), M, n = 9
0.1% Pb, M, n = 9
Day 270
Oral, drinking water
0.34 ± 0.23 pg/dLfor
0.0%, 7.21 ± 1.27 pg/dL
for 0.1%
transepithelial electrical
potential difference (PD),
changes in the
transepithelial electrical
potential difference
during mechanical
stimulation (dPD),
transepithelial electrical
resistance (R)
Reddvetal. (2018)
Rat (Sprague Dawley),
Control Diet (CD), M,
n = 10
Control Diet, F, n = 10
Iron Deficient (ID), M,
n = 10
Iron Deficient, F, n = 10
Control Diet + Pb, M,
n = 10
Control Diet + Pb, F,
n = 10
Iron Deficient + Pb, M,
n = 10
Iron Deficient, F, n = 10
Microbiome Counts at
Week 0, 4, 8, 10, 12
BLL at End of Week 8
Oral, gavage
2.3 ± 1.16 |jg/dL - CD, M
19.3 ±6.23 pg/dL-
CD + Pb, M
2.5 ± 0.89 |jg/dL - ID, M
47.5 ± 3.78 pg/dL-
ID + Pb, M
1.9 ± 0.81 |jg/dL - CD, F
13.5 ± 3.52 pg/dL-
CD + Pb, F
1.5 ± 0.31 |jg/dL - ID, F
29.80 ± 8.30 pg/dL-
ID + Pb, F
Fecal Lactobacilli
(Counts), Fecal E. Coli
(Counts), Fecal Yeast
(Counts)
BLL = blood lead level; CD = control diet; dPD = transepithelial electrical potential difference during mechanical stimulation; F = female; ID = iron deficient; M = male;
PD = transepithelial electrical potential difference; R = resistance
9-102
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Table 9-9
Epidemiologic studies of exposure to Pb and endocrine effects
Reference and
Study Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tChen etal. (2013)
United States
2007-2008
Cross-sectional
NHANES
n = 5,418
Adolescents and adults in
the general U.S.
population who had no
reported thyroid diseases,
thyroid medications,
pregnancy, and sex
steroid medications.
Blood Pb
Blood Pb was
measured in venous
whole blood using
GFAAS
Age at measurement:
>12 yr old
Mean: 0.93 |jg/dL
Max: 9.20 |jg/dL
TSH, thyroglobulin (Tg),
and thyroid hormones (T3,
FT3, T4, FT4)
TSH and thyroid
hormones measured in
serum using the Beckman
Immunoassay System.
Age at outcome:
>12 yr old
Age, sex, race/ethnicity,
creatinine-adjusted
urinary iodine, BMI Z-
score, and serum
cotinine level
Change in T4 ([jg/dL)b
Adolescents (12-19 yr old)
(-0.02, 0.04)
Adults (>19 yr old)
-0.01 (-0.02, 0.01)
Change in FT4 (ng/dL)b
Adolescents (12-19 yr old)
(-0.01, 0.04)
Adults (>19 yr old)
0.01 (-0.01, 0.02)
Change in T3 (ng/dL)b
Adolescents (12-19 yr old)
(-0.01, 0.04)
Adults (>19 yr old)
-0.0004 (-0.02, 0.02)
Change in FT3 (pg/ml_)b
Adolescents (12-19 yr old)
(-0.002, 0.04)
Adults (>19 yr old)
0.01 (-0.001, 0.02)
Change in TSH ([jlU/mL)b
Adolescents (12-19 yr old)
-0.05 (-0.18, 0.07)
Adults (>19 yr old)
-0.01 (-0.06, 0.04)
Change in Tg (ng/ml_)b
Adolescents (12-19 yr old)
0.05 (-0.13, 0.24)
Adults (>19yrold)
0.01 (-0.03, 0.06)
9-103
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Reference and
Study Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tKrieq (2019)
United States
1988-1994
Cross-sectional
NHANES III
n = 16,573
General population,
>20 yr old
Blood Pb
Blood Pb was
measured in venous
whole blood using
AAS
Age at measurement:
>20 yr old
Mean: 3.55 |jg/dL
(SE = 0.10)
TSH and T4
TSH and thyroid
hormones measured in
serum using the Beckman
Immunoassay System.
Age at outcome:
>20 yr old
Linear regression model
adjusted for race-
ethnicity, sex, age,
session, BMI, pregnant,
menopause, hormone
pill use, vaginal cream
use, hormone patch
use, urinary creatinine
Change in TSH (%)
-1.2 (-5.6, 3.3)
Change in T4 (%)
-38.9 (-51.3, -23.4)
Change in Logio-TSH
(HU/mL)b
Male
0.01 (-0.04, 0.05)
Female (Not pregnant)
-0.04 (-0.08, 0.01)
Female (Pregnant)
-0.03 (-0.26, 0.20)
Change in Logio-T4 ([jg/dL)b
Male
-0.15 (-0.48, 0.18)
Female (Not pregnant)
-0.52 (-0.83, -0.21)
Female (Pregnant)
-2.01 (-3.09, -0.93)
9-104
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Reference and
Study Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tMendv et al. (2013) NHANES
n = 4,652
United States
2007-2008
Cross-sectional
General population >20 yr
old, excluding pregnant
women, individuals with a
history of thyroid disease,
or under treatment for
thyroid dysfunction
Blood Pb
Blood Pb was
measured in venous
whole blood using
GFAAS
Age at measurement:
>20 yr old
Mean (SD):
1.52 ± 1.20 |jg/dL
Max: 33.12 pg/dL
TSH and thyroid
hormones (T3, FT3, T4,
FT4)
TSH and thyroid
hormones measured in
serum using the Beckman
Immunoassay System
Age at outcome:
>20 yr old
Age, gender,
race/ethnicity, smoking,
alcohol consumption,
BMI, physical activity,
iodine intake,
medications, and bone
mineral density
Change in T3 (ng/dL)
-0.774 (-2.269, 0.722)
Change in FT3 (pg/mL)
0.015 (-0.007, 0.037)
Change in T4 ((jg/dL)
-0.162 (-0.321, -0.004)
Change in FT4 (ng/mL)
(-0.011, 0.011)
Change in TSH (mlll/mL)
0.015 (-0.088, 0.118)
tChristensen (2012)
United States
2007-2008
Cross-sectional
NHANES
n = 1,587
General population,
>20 yr old, excluding
individuals with thyroid
disease or cancer, or
were taking thyroid
medications
Blood Pb
Blood Pb was
measured in venous
whole blood using
GFAAS
Age at measurement:
>20 yr old
Median: 1.3 pg/dL
75th: 2.1 pg/dL
TSH and thyroid
hormones (T3, T4)
TSH and thyroid
hormones measured in
serum using the Beckman
Immunoassay System.
Age at outcome:
>20 yr old
Age, sex, race, BMI,
serum lipids, serum
cotinine, pregnancy and
menopausal status, and
use of selected
medications
Change in ln(T3) (ng/dL)
0.004 (-0.016, 0.023)
Change in ln(FT3) (pg/mL)
0.008 (-0.002, 0.017)
Change in ln(T4) (pg/dL)
-0.018 (-0.036, 0)
Change in ln(FT4) (pg/mL)
-0.001 (-0.018, 0.015)
Change in In(TSH) (mlU/L)
0.027 (-0.031, 0.085)
9-105
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Reference and
Study Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tLuo and Hendrvx
(2014)
United States
2007-2010
Cross-sectional
NHANES
n = 6,231
General population >20 yr
old, excluding pregnant
women, individuals with
history of thyroid disease,
or missing data.
Blood Pb
Blood Pb was
measured in venous
whole blood using
GFAAS
Age at measurement:
>20 yr old
Mean: 1.82 |jg/dL
Max: 33.10 |jg/dL
TSH, thyroglobulin (Tg),
and thyroid hormones (T3,
FT3, T4, FT4)
TSH and thyroid
hormones measured in
serum using the Beckman
Immunoassay System.
Age at outcome:
>20 yr old
Adjusted for age, sex,
race and ethnicity,
serum cotinine, BMI,
and creatinine-adjusted
urinary iodine
Change in T3 across tertiles
(ng/dL)b
T1: Reference
T2: 1.02 (-0.90, 2.94)
T3: 0.69 (-2.37, 3.76)
Women Only
T1: Reference
T2: -0.36 (-3.72, 3.00)
T3: 0.61 (-5.02, 6.23)
Men Only
T1
T2
T3
Reference
1.96 (-0.98, 4.91)
0.69 (-2.59, 3.97)
Change in FT3 across
tertiles (pg/ml_)b:
T1: Reference
T2: 0.03 (0.001, 0.07)
T3: 0.04 (0.01, 0.08)
Women Only
T1
T2
T3
Reference
0.02 (-0.04, 0.08)
0.03 (-0.04, 0.11)
Men Only
T1
T2
T3
Reference
0.03 (-0.01, 0.07)
0.05 (0.01, 0.09)
Change in T4 across tertiles
(ng/dL)b:
T1: Reference
T2: 0.01 (-0.16, 0.14)
9-106
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Reference and
Study Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
T3: -0.09 (-0.28, 0.11
Women Only
T1
T2
T3
Reference
0.12 (-0.10, 0.35)
0.02 (-0.29, 0.33)
Men Only
T1
T2
T3
Reference
-0.14 (-0.35, 0.08)
-0.20 (-0.40, 0.01)
Change in FT4 across
tertiles (ng/dL)b:
T1: Reference
T2: 0.007 (-0.01, 0.02)
T3: 0.002 (-0.01, 0.01)
Women Only
T1: Reference
T2: 0.02 (0.01, 0.04)
T3: 0.02 (-0.003, 0.04)
Men Only
T1: Reference
T2: -0.02 (-0.03, 0.005)
T3: -0.01 (-0.04, 0.008)
Change in Log-Tg across
tertiles (ng/ml_)b:
T1: Reference
T2: 0.04 (-0.04, 0.13)
T3: 0.02 (-0.07, 0.12)
Women Only
T1: Reference
T2: 0.08 (-0.03, 0.19)
9-107
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Reference and
Study Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
T3: -0.06 (-0.19, 0.08)
Men Only
T1: Reference
T2: -0.001 (-0.13, 0.17)
T3: 0.05 (-0.08, 0.17)
Change in Log-TSH across
tertiles (ulll/mL)b:
T1: Reference
T2: 0.01 (-0.05, 0.07)
T3: 0.02 (-0.06, 0.09)
Women Only
T1: Reference
T2: 0.05 (-0.06, 0.16)
T3: 0.02 (-0.09, 0.14)
Men Only
T1: Reference
T2: -0.04 (-0.13, 0.06)
T3: -0.02 (-0.11, 0.07)
9-108
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Reference and
Study Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tNie etal. (2017)
Shanghai and 7
provinces
China
2014
Cross-sectional
SPECT-China study
n = 5,628
Residents of these
regions are 99.5% Han
Chinese.
Exclusion criteria
included age under 18 yr
old, less than 6 mo spent
at current residence, and
severe communication
problems or acute illness
(thyroid resection or
iodine-131 therapy,
malignant tumor,
subacute thyroiditis, liver
cirrhosis)
Blood Pb
Whole blood
measured using AAS
Age at measurement:
18-93 yr old
Median:
Men: 44.00 |jg/L
Women: 37.87 |jg/L
Mean:
Men:
29.00 ± 62.18 |jg/L
Women:
25.03 ± 54.61 |jg/L
TSH, thyroid hormones
(T3, T4), thyroid
peroxidase antibody
(TPOAb) and thyroglobulin
antibody (TGAb)
Thyroid dysfunction and
subclinical thyroid
dysfunction were
measured by
immunochemiluminometric
assays
Age at outcome:
18-93 yr old
Linear and logistic
regression model
adjusted for age, BMI
smoking status (men
only) and drinking status Women
1.41 (0.00, 2.84)
Change in TPOAb (%)
Men
0.50 (-0.80, 1.82)
Change in TGAb (%)
Men
-0.60 (-1.88, 0.70)
Women
0.20 (-1.09, 1.51)
Change in TSH (%)
Men
-0.40 (-1.29, 0.40)
Women
1.11 (0.30, 1.82)
tKahn etal. (2014)
Pristina and Mitrovica
Yugoslavia
1985-1986
Cross-sectional
Yugoslavia Prospective
Study of Environmental
Pb Exposure
n = 291
Pregnant women in
second trimester, major
central nervous system
defects, multiple births,
and residence >10 km
from clinic
Blood Pb
Whole blood samples
taken in Yugoslavia
and transported on
wet ice to Columbia
University. Blood Pb
measured using
GFAAS.
Age at measurement:
16-41 yr old
Mean |jg/dL (SD):
Pristina: 5.57 (2.01)
Mitrovica: 20.00 (6.99)
TSH, thyroid hormones
(FT4), and thyroid
peroxidase antibodies
(TPOAb)
FT4 and TPOAb
were measured by a
radioimmunoassay
procedure. TSH was
measured using an
immunoradiometric assay
Age at outcome:
16-41 yr old
Logistic regression
model adjusted for:
FT4: height, ethnicity,
BMI, fetal gestational
age, maternal
education, adults per
room; TSH: hemoglobin,
ethnicity, BMI, fetal
gestational age,
maternal age; TPOAb:
ethnicity, fetal
gestational age,
maternal age, adults per
room.
Change in FT4 (ng/dL)b
-0.074 (-0.10, -0.046)
Change in Log-TSH
(HlU/mL)b
0.026 (-0.065, 0.12)
Change in Log-TPOAb
(IU/mL)b
0.31 (0.17, 0.46)
ORb
TPOAb > vs. <10 lU/mL
2.41 (1.53, 3.82)
9-109
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Reference and
Study Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tSouza-Talarico et al.
(2017)
Sao Paulo
Brazil
Cross-sectional
N = 126
105 women and 21 men
ages 50-82 yr old with a
mean of 9.8 (±4.5) yr of
education
Blood Pb
Fasting blood Pb was
measured using ICP-
MS
Age at measurement:
50-82 yr old
Median: 2.1 |jg/dL
(SD: ±0.9)
Max: 6.1 |jg/dL
Cortisol concentration and
allostatic load
Six neuroendocrine,
metabolic, and
anthropometric biomarkers
were analyzed, and values
were transformed into an
AL index using clinical
reference cut-offs. Salivary
samples were collected at
home over 2 d at
awakening, 30-min after
waking, afternoon, and
evening periods to
determine Cortisol levels.
Age, gender, time of
awakening,
socioeconomic status
(SES), GDS, and PSS
scores
Change in CAR (pg/dL min)b
0.791 (0.672, 1.073)
Change in total AUC (jjg/dL
hr)b
0.889 (0.829, 0.953)
Age at outcome:
50-82 yr old
9-110
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Reference and
Study Design
Study Population
Exposure
Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tNqueta et al. (2018) Study of Genetics, Stress Blood Pb
Montreal
Canada
2004-2006
Cross-sectional
and Cognitive
Development
n = 65
75% of participants were
women, 95% were
Caucasian, 90% were
current smokers
Blood Pb levels were
determined using
inductively coupled
plasma mass
spectroscopy
Age at measurement:
50-67 yr old
Median: 2.48 |jg/dL
Mean: 2.41 |jg/dL
(SD = 0.15)
Diurnal basal Cortisol
levels and acute Cortisol
responsivity
Basal Cortisol:
Participants were
instructed to collect saliva
five times per day during
three consecutive
weekdays: upon
awakening, 30 min after
awakening, at 2:00 p.m.,
at 4:00 p.m., and at
bedtime
Linear model adjusted
for age, gender, waist-
hip ratio, smoking status
and income levels.
Change in basal Cortisol
levels (|jg/dL)
-0.01 (-0.05, 0.02)
Change in reactive Cortisol
levels (|jg/dL)
-0.01 (-0.03 0.01)
Stress reactivity: A total of
nine saliva samples were
collected for measurement
of salivary Cortisol: two
baseline samples, one
postanticipatory, and six
post-TSST tasks: one after
15 min and then five
sampled every 10 min
Age at outcome:
50-67 yr old
AAS = atomic absorption spectrometry; BMI = body mass index; CAR = Cortisol awakening response; CI = confidence interval; d = day(s); GDS = Gesell Developmental Schedules;
GFAAS = graphite furnace atomic absorption spectrometry; FT3 = free triiodothyronine; FT4 = free thyroxine; ICP-MS = inductively coupled plasma mass spectrometry;
NHANES = National Health and Nutrition Examination Survey; Pb = lead; PSS = perceived stress score;SD = standard deviation; SE = standard error; SES = socioeconomic status;
SPECT = single photon emission computed tomography; Tg = thyroglobulin; T# = fertile #; TGAb = thyroglobulin antibody; TPOAb = thyroid peroxidade antibody;
TSH = thyroid-stimulating hormone; yr = year(s)
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bEffect estimate unable to be standardized due to insufficient distribution information.
fStudies published since the 2013 Integrated Science Assessment for Pb.
9-111
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Table 9-10
Animal toxicological studies of exposure to Pb and endocrine effects
Study
Species (Stock/Strain), n, Sex Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (jjg/dL)b Corticosterone Levels
Rossi-George et al.
(2011)
Rat (Long-Evans)
Control (untreated), M/F, n = 10
dams
50 ppm, M/F, n = 9 dams
150 ppm, M/F, n = 11 dams
GD-61 to Dams were dosed
PND 304 starting 2 mo prior to
mating through
lactation. Pups were
weaned on PND 21
and continued on the
regimen of their dam
until euthanasia post-
testing at approximately
10 mo of age.
0.979 |jg/dL for 0 ppm,
19.091 |jg/dL for 50 ppm,
35.245 |jg/dL for 150 ppm ¦
PND 21 Females
I.469 |jg/dL for 0 ppm,
II.259 |jg/dLfor50 ppm,
25.699 |jg/dL for 150 ppm ¦
PND 61 Females
I.713 |jg/dL for 0 ppm,
II.993 |jg/dLfor50 ppm,
29.615 |jg/dL for 150 ppm ¦
PND 304 Females
Adrenal Weight,
Corticosterone Levels
1.935 |jg/dL for 0 ppm,
19.597 |jg/dL for 50 ppm,
31.935 |jg/dL for 150 ppm-
PND 21 Males
2.177 |jg/dL for 0 ppm,
12.581 |jg/dL for 50 ppm,
26.855 |jg/dL for 150 ppm -
PND 61 Males
1.694 |jg/dL for 0 ppm,
15.968 |jg/dL for 50 ppm,
29.274 |jg/dL for 150 ppm -
PND 304 Males
9-112
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Study
Species (Stock/Strain), n, Sex
Timing of
Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (jjg/dL)b Corticosterone Levels
Graham et al. Rat (Sprague Dawley)
(2011) Control (vehicle), M/F,
n = 12-18 (6-8/6-8)
1 mg/kg Pb, M/F, n = 12-18
(6-8/6-8)
10 mg/kg Pb, M/F, n = 12-18
(6-8/6-8)
PND 4 to Rats were gavaged
PND 28 every other day from
P4 until P28.
0.267 |jg/dL for 0 mg/kg,
3.27 |jg/dL for 1 mg/kg
12.5 |jg/dL for 10 mg/kg ¦
PND 29
Adrenal Weight,
Corticosterone Levels
Corv-Slechta et al.
(2013)
Mouse (C57BL.6) GD-61 to
Control (untreated), M, n = 8-17 PND 365
Control (untreated), F, n = 8-13
100 ppm Pb, M, n = 8-17
100 ppm Pb, F, n = 8-13
Dams were exposed
starting 2 mo prior to
mating. Offspring were
continued on the same
exposure as their dams
until the end of the
experiment at 12 mo of
age.
0.34 |jg/dL for 0 ppm Fl
males
0.11 |jg/dL for 0 ppm FS
males
0.34 |jg/dL for 0 ppm Fl
females
0.16 |jg/dL for 0 ppm FS
females
6.94 |jg/dL for 100 ppm Fl
males
6.16 |jg/dL for 100 ppm FS
males
9.38 |jg/dL for 100 ppm Fl
females
7.07 |jg/dL for 100 ppm FS
females
Adrenal Weight,
Corticosterone Levels
Amos-Kroohs et al.
(2016)
Rat (Sprague Dawley)
Control (vehicle, see notes),
M/F, n = 16 (8/8)
1 mg/kg Pb, M/F, n = 16 (8/8)
10 mg/kg Pb, M/F, n = 16 (8/8)
P4 until P10,
18, or 28.
Rats were gavaged
every other day from
PND4 until PND10, 18,
or 28.
1.24 |jg/dL for 0 mg/kg Pb
2.79 |jg/dL for 1 mg/kg Pb
9.07 |jg/dL for 10 mg/kg Pb
Corticosterone Levels
9-113
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Study
Species (Stock/Strain), n, Sex
Timing of
Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (jjg/dL)b Corticosterone Levels
Sobolewski et al.
(2020)
Mouse (C57BL.6)
F0 Control (assume untreated),
F, n = 10
100 ppm Pb, F, n = 10
20 females were in control and
20 received Pb but these groups
were further divided, and some
received prenatal stress and
others did not.
GD -61 to Exposure started 2 mo
PND 21 prior to mating and
continued through
PND 21 (weaning) of
the F1.
F3 was technically not
directly exposed.
F1 0.0 |jg/dL for Control
12.5 |jg/dL for 100 ppm -
PND 6-7
F3 0.0 |jg/dL for Control
Corticosterone Levels
F# = filial generation; F = female; GD = gestational day; M = male; mo = month(s); Pb = lead; PND = postnatal day.
9-114
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Table 9-11
Epidemiologic studies of exposure to Pb and musculoskeletal effects
Reference and
Study Design
Study Population Exposure Assessment Outcome Confounders
Effect Estimates and
95% Clsa
Osteoporosis and Bone Mineral Density
tChoetal. (2012)
South Korea
2008
Cross-Sectional
KNHANES
n = 481
Postmenopausal women
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
Mean (SD):
Q1: 64.03 (8.52) yr
Q2 and Q3: NR
Q4: 61.78 (8.62) yr
Median: 2.32 |jg/dL
25th: 1.83 pg/dL
75th: 2.88 pg/dL
Osteoporosis
BMD measured in hip,
neck, and spine using X-
ray absorptiometry.
Osteoporosis defined as
T-score <2.5 at any of
the measurement sites
Age at outcome is the
same as the age at
exposure assessment
Age, BMI, alcohol
intake, cigarette
smoking, exercise,
use of oral
contraceptive pill,
hormone therapy,
caloric intake,
calcium intake, fish
consumption, and
vitamin D level
OR Osteoporosis
Prevalence
Q1:
Ref.
Q2:
1.41
(0.75,
2.67)
Q3:
1.34
(0.70,
2.56)
Q4:
1.50
(0.79,
2.86)
tWanq et al. (2019) NHANES
n = 1859
Blood, Urine
United States
2013-2014
Cross-sectional
Blood Pb measured in whole
General population; >40 yr blood using ICP-MS
old
Age at measurement:
>40 yr
Mean: 1.24 pg/dL
75th: 1.81 pg/dL
BMD and fracture risk
BMD measured via DXA
scan; Fracture risk
measured via Fracture
Risk Assessment score -
a composite index of
fracture risk factors
Age at outcome:
>40 yr
Age, race/ethnicity,
BMI, serum 25(OH)D
level, smoking,
drinking, treatment
for osteoporosis, and .. .
r i • ivfdlQS
use of prednisone
Change in BMD (g/cm2)
Femur
-0.01 (-0.03, 0.01)
Premenopausal Women
-0.06 (-0.08, -0.03)
Menopausal Women
0.01 (-0.01, 0.03)
Spine
Males
0.01 (-0.01, 0.03)
9-115
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Premenopausal Women
-0.05 (-0.08, -0.02)
Menopausal Women
0.02 (-0.01, 0.04)
Change in 10-yr
Fracture Risk Score
Hip
0.45 (0.28, 0.62)
Major
1.22 (0.68, 1.77)
tLee and Kim
(2012)
South Korea
2008-2009
Cross-Sectional
KNHANES
n = 832
Women ages >40 yr
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
Mean (SD): 56.1 (10.4) yr
GM: 2.182 pg/dL
BMD
BMD in the femoral neck,
trochanter,
intertrochanter, Ward
triangle, total femur, and
lumbar 1-4. Measured
using DXA
Age at outcome:
Mean (SD): 56.1 (10.4)
yr
Residence area,
obesity, educational
level, smoking status,
drinking status,
number of
pregnancies,
hormone treatment,
contraceptive oral pill
and daily calcium
intake for pre- and
postmenopausal, and
time since
menopause for
postmenopausal
Change in BMD (g/cm2)
Premenopausal Women
Total Femur
-0.15 (-0.33, 0.03)
Trochanter
-0.18 (-0.41, 0.05)
Intertrochanter
-0.11 (-0.25, 0.03)
Femoral Neck
-0.11 (-0.28, 0.07)
Ward's Triangle
-0.11 (-0.26, 0.03)
9-116
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Lumbar 1-4
-0.09 (-0.24, 0.06)
Menopausal Women
Total Femur
-0.28 (-0.45, -0.11)
Trochanter
-0.30 (-0.55, -0.06)
Intertrochanter
-0.22 (-0.35, -0.08)
Femoral Neck
-0.21 (-0.39, -0.02)
Ward's Triangle
-0.13 (-0.29, 0.03)
Lumbar 1-4
-0.17 (-0.31, -0.04)
tPollack et al.
(2013)
BioCycle Study
n = 248
Western New York Premenopausal women
United States ages 18-44 yr
2005-2007
Cross-Sectional
Blood
Bone mineral density
Blood Pb measured in venous BMD in the hip, spine,
whole blood using ICP-MS
Age at measurement:
Mean (SD): 27.4 (8.2) yr
Mean: 1.03 |jg/dL
wrist, and whole body
(g/cm2) measured via
DXA
Age at outcome:
Mean (SD): 27.4 (8.2) yr
Age, BMI, race, Change in BMD (g/cm2)
parity, caloric intake,
and age at menarche „ ,
a Whole Body
-0.004 (-0.03, 0.021)
Total Hip
-0.002 (-0.035, 0.031)
Lumbar Spine
9-117
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
-0.016 (-0.048, 0.016)
Wrist
0.001 (-0.012, 0.014)
tLi et al. (2020b)
Sichuan Province
China
Cross-sectional
n = 799
Study area included two
rural towns, one with a
history of heavy metal
contamination. Generally
healthy adults ages 40-
75 yr old who lived in
study area for >15 yr and
subsisted on rice and
vegetables grown in study
area.
Blood, Urine
Blood Pb measured in venous
whole blood using ICP-MS
Age at measurement:
40-75 yr
Median 3.4 |jg/dL
75th: 4.7 pg/dL
BMD
Osteoporosis (BMD T-
score <2.0); BMD
measured via X-ray
absorptiometry
Age at outcome:
40-75 yr
Age, BMI, and
smoking status
OR Osteoporosis
Prevalence (>3.4 pg/dL
vs. <3.4 [jg/dL blood Pb)
Males
0.6 (0.24, 1.49)
Females
1.33 (0.61, 2.88)
Non-Smoking Females
0.94 (0.4, 2.21)
tLimetal. (2016)
South Korea
2008-2011
Cross-Sectional
KNHANES Blood
n = 2429
Blood Pb measured in venous
General population; >18 yr whole blood using GFAAS
old
Age at measurement:
>18 yr
Median: 2.22 pg/dL
25th: 1.66 pg/dL
75th: 2.93 pg/dL
BMD (osteoporosis and
osteopenia)
Osteopenia (BMD T-
score <—1.0) and
Osteoporosis (BMD T-
score <-2.5)
Age at outcome:
>18 yr
Age, sex, smoking
status, alcohol
consumption,
geographic region,
education level,
occupation, and
family income
ORs for Osteoporosis or
Osteopenia prevalence
across blood Pb
quartiles
Q1
Q2
Q3
Q4
Ref.
1.08 (0.85, 1.37)
1.18 (0.91, 1.53)
1.49 (1.12, 1.98)
9-118
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tLee and Park
(2018)
Ansung and Ansan
South Korea
2001-2002
Cross-Sectional
Korean Association
Resource (KARE) Cohort
n = 443
Adults aged 40-65 yr from
two South Korean
communities, on rural
(Ansung) and one urban
(Ansan)
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
40-65 yr
GM: 4.44 pg/dL
BMD
Age, sex, geographic Change in BMD T-score
region, income, and
BMD (T-score) physical activity
measured via ultrasound
Age at outcome:
40-65 yr
All
-0-0.26 (-0.45, -0.07)
Ever Smokers
-0.47 (-0.85, -0.09)
Current Smokers
-0.60 (-1.02, -0.17)
Never Smokers
-0.15 (-0.37, 0.07)
Osteoarthritis
tPark and Choi
(2019)
South Korea
4 Years (2010—
2012)
Cross-sectional
KNHANES
n = 884
Women, >55 yr old
Blood
BLL measured in venous whole
blood using GFAAS
Age at measurement:
Mean: 62.9 yr
Median: 2.22 pg/dL
Max: 7.84 pg/dL
Osteoarthritis
Radiographic and
symptomatic
osteoarthritis.
Radiographic OA (rOA)
assessed in the hip,
knee, and spine using
the Kellgren-Lawrence
grading system.
Symptomatic OA (sxOA)
assessed using a
combination of
radiographic evidence
and self-reported
symptoms
Age at outcome:
Mean: 62.9 yr
Age, smoking status,
alcohol use, physical
activity, education,
occupation, income,
diabetes,
hypertension, and
BMI
ORs for Osteoarthritis
prevalence per In-unit
increase in blood Pb
(Hg/dL)
rOA Knee
1.77 (1.17, 2.67)
sxOA Knee
1.50 (0.90, 2.53)
rOA Back
1.05 (0.70, 1.59)
sxOA Back
0.68 (0.39, 1.18)
9-119
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tNelson et al.
(2011a)
Johnston County,
N.C.
United States
2003-2004 and
2006-2008
Cross-Sectional
Johnston County
Osteoarthritis Project
n = 668
African American and
white adults ages >45 yr
old
Blood
Blood Pb measured in venous
whole blood using ICP-MS
Age at measurement
Mean (SD):
Females: 62.4 (9.4) yr
Males: 64.5 (10.8) yr
Median:
Females: 1.9 |jg/dL
Males: 2.2 |jg/dL
Max:
Females: 25.4 |jg/dL
Males: 25.1 |jg/dL
Osteoarthritis
Urine and serum
biomarkers of joint tissue
metabolism
Age at outcome:
Mean (SD):
Females: 62.4 (9.4) yr
Males: 64.5 (10.8) yr
Age, race, BMI, and
smoking status
% Change in urine and
serum biomarkers of
joint tissue metabolism
Males
uNTX-l
1.2% (-1.0, 3.4%)
UCTX-II
1.4% (-0.6, 3.4%)
COMP
1.6% (-0.1, 3.2%)
C2C
0.0% (-1.0, 1.0%)
CPU
-0.2% (-1.4, 1.0%)
C2C.CPII
0.0% (-1.4, 1.4%)
HA
0.2% (-2.5, 3.0%)
Females
uNTX-l
7.7% (3.9, 11.7%)
9-120
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
UCTX-II
5.1% (0.8, 9.5%)
COMP
-0.8% (-2.8, 1.2%)
C2C
0.0% (-1.6, 1.6%)
CPII
1.7% (-0.6, 4.1%)
C2C.CPII
-1.2% (-3.5, 1.1%)
HA
-0.8% (-6.6, 5.3%)
9-121
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tNelson et al.
(2011b)
Johnston County,
N.C.
United States
2003-2004 and
2006-2008
Cross-Sectional
Johnston County
Osteoarthritis Project
n = 1635
African American and
white adults ages >45 yr
old
Blood
Blood Pb measured in venous
whole blood using ICP-MS
Age at measurement:
Mean (SD): 65.3 (11.0) yr
Mean: 2.4 |jg/dL
Osteoarthritis
Radiographic and
symptomatic
osteoarthritis.
Radiographic OA (rOA)
assessed in the knee
using the Kellgren-
Lawrence grading
system. Symptomatic OA
(sxOA) assessed using a
combination of
radiographic evidence
and self-reported
symptoms
Age at outcome:
Mean (SD): 65.3 (11.0)
yr
Age, sex, race,
ethnicity, BMI,
current smoking, and
current drinking
ORs for Prevalent
Osteoarthritis of the
Knee
rOA
1.10 (1.00,
sxOA
1.08 (0.96,
1.20)
1.20)
Oral Health - Adults
tWon et al. (2013)
South Korea
2009
Cross-Sectional
KNHANES
n = 1966
General population; >19 yr
old
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
>19 yr
Mean NR
T1: <1.73 |jg/dL
T2: 1.73-3.04 pg/dL
T3: >3.04 pg/dL
Periodontal disease
Community Periodontal
Index (code >3,
corresponding to pockets
>3.5 mm)
Age at outcome:
>19 yr
Age, sex, family
income, education
level, use of floss,
use of interproximal
toothbrush, alcohol
consumption,
smoking status, ETS
in workplace,
diabetes,
hypertension, and
oral health status
ORs for Prevalent
Periodontal Disease
across blood Pb tertiles
T1
T2
T3
Ref.
1.37 (0.97,
1.31 (0.88,
1.93)
1.96)
9-122
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
1-Han etal. (2013)
South Korea
2008-2010
Cross-Sectional
KNHANES
n = 4716
General population; >19 yr
old
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
>19 yr
GM:
Periodontitis: 2.60 |jg/dL
No periodontitis: 2.12 |jg/dL
Periodontal disease
Community Periodontal
Index (code >3,
corresponding to pockets
>3.5 mm)
Age at outcome:
>19 yr
Age, gender, income,
education, frequency
of daily
toothbrushing,
regular dental check-
up, smoking, alcohol
consumption,
physical activity,
fasting plasma
glucose, BMI, white
blood cell count and
urine cotinine
concentration.
ORs for Prevalent
Periodontal Disease
across blood Pb
quintiles
Q1 (<1.59 |jg/dL)
Ref.
Q2 (1.59-2.05 |jg/dL)
1.36 (1.00, 1.85)
Q3 (2.05-2.52 pg/dL)
1.3 (0.96, 1.76)
Q4 (2.52-3.57 pg/dL)
1.55 (1.13, 2.13)
Q5 (>3.17 pg/dL)
1.6 (1.15, 2.22)
tKim and Lee
(2013)
South Korea
2008-2009
Cross-Sectional
KNHANES Blood
n = 3996
Blood Pb measured in venous
General population; >20 yr whole blood using GFAAS
old
Age at measurement:
>20 yr
GM: 2.31 pg/dL
Periodontal Disease
Community Periodontal
Index (code >3,
corresponding to pockets
>3.5 mm)
Age at outcome:
>20 yr
Age, body mass
index (BMI),
residence area,
education level,
household income,
smoking and drinking
status, hemoglobin,
glucose, use of floss
or interproximal
toothbrush, decayed,
missing, or filled
permanent teeth
(DMFT), and active
caries
ORs for Prevalent
Periodontal Disease
across blood Pb
quintiles
Males
1.854 (1.265, 2.717)
Males (adjusted for Hg,
Cd)
1.699 (1.154, 2.502)
Females
1.301 (0.883, 1.917)
Females (w/ Hg and Cd)
9-123
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
1.242 (0.833, 1.851)
Oral Health - Children and Adolescents
tWu etal. (2019)
Mexico City
Mexico
Initial Recruitment:
1997-2005; Follow-
up: 2008-2013
Cohort
Early Life Exposure in
Mexico to Environmental
Toxicants (ELEMENT)
n = 173 to 386 (depending
on exposure metric)
Mother/child pairs
recruited from 2 public
hospitals serving low-to
moderate-income
populations
Blood
Maternal and child blood Pb
measured in venous whole
blood using GFAAS. Maternal
bone Pb measured using K-XRF
Age at measurement:
Maternal BLL:
1st, 2nd, and 3rd trimester
Child BLL:
1, 2, 3, and 4 yr, and in
adolescence (10 to 18 yr)
Maternal bone:
Postnatally
Mean (males, females):
1st trimester: 6.06, 6.36 |jg/dL
2nd trimester: 5.24, 5.25 |jg/dL
3rd trimester: 5.67, 5.73 |jg/dL
Childhood: 15.48, 15.18 pg/dL
Adolescence: 3.60, 3.34 pg/dL
Maternal tibia: 8.64, 9.68 pg/g
Maternal patella: 7.18, 8.64 pg/g
Dental caries
Teeth evaluated by
trained examiners who
assigned decayed,
missing, filled tooth
(DMFT) scores
Age at outcome:
Adolescence (10 to
18 yr)
Sex, cohort, mother's Rate Ratio of Decayed,
education, sugar Missing, and Filled
sweetened Teeth per In-unit
beverages intake increase in blood or
bone Pb
1st Trimester BLL
1.07 (0.90, 1.27)
2nd Trimester BLL
1.12 (0.94, 1.32)
3rd Trimester BLL
1.17 (0.99, 1.37)
Childhood BLL
1.14 (0.94, 1.38)
Adolescent BLL
0.97 (0.81, 1.16)
Maternal Patella Pb
0.95 (0.88, 1.03)
Maternal Tibia Pb
0.98 (0.88, 1.08)
9-124
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tYepes et al. (2020) Early Life Exposure in
Mexico City
Mexico
Initial Recruitment:
1997-2005; Follow-
up: 2008-2013
Cohort
Mexico to Environmental
Toxicants (ELEMENT)
n = 490
Mexican children recruited
from 2 public hospitals
serving low-to moderate-
income populations
Blood
Child blood Pb measured in
venous whole blood using
GFAAS.
Child BLL:
Average of measurements at
ages 1, 2, 3, and 4 yr
Mean (SD): 4.83 (2.2) pg/dL
Dental caries
Teeth evaluated by
trained examiners who
assigned decayed,
missing, filled tooth
(DMFT) scores
Age at outcome:
Adolescence (9 to 17 yr)
Age, sex, BMI,
sugar intake (g/day),
water intake (ml/day),
amount of beverages
with sugar per day
(mL/day), amount of
beverages without
sugar per day
(mL/day), amount of
toothpaste used
regularly from birth to
2 yr, amount of
toothpaste used from
2 to 4 yr, amount of
toothpaste used from
4 to 6 yr, and amount
of current toothpaste
use
D1MFS Beta (increment
not reported)
Mean BLL
0.03 (-0.03,
Peak BLL
0.01 (-0.01,
0.09)
0.04)
tKimetal. (2017)
Seoul, Daegu,
Cheonan, and
Busan
South Korea
2005-2010
Cross-sectional
The Children's Health and
Environmental Research
(CHEER) group
n = 1,565 (children w/
permanent teeth) and
1,241 (children w/
deciduous teeth)
School-aged children from
urban, rural, and
industrialized areas with
BLLs <5 pg/dL
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
"School-aged"
GM: 1.53 pg/dL
Dental caries
DMFS sum by trained
dental hygienists
Age at outcome:
"School-aged"
Sex, age
(categorical),
household income
(categorical), and
urinary cotinine level
(categorical)
PR for Decayed and
Filled Surfaces
Deciduous Teeth
Decayed Surfaces
1.16 (0.91, 1.49)
Filled Surfaces
1.11 (0.98, 1.25)
DMFS
1.14 (1.02, 1.27)
Permanent Teeth
Decayed Surfaces
0.69 (0.45, 1.07)
9-125
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Filled Surfaces
0.87 (0.73, 1.04)
DMFS
0.83 (0.69, 0.99)
tWiener et al.
(2015)
United States
1988-1994
Cross-Sectional
NHANES III
n = 3127
General population; 2 to
6 yr old
Blood
Blood Pb measured in venous
whole blood using GFAAS
Age at measurement:
2 to 6 yr
Mean NR
28.2% <2 |jg/dL;
48.3% 2 to <5 |jg/dL;
18.4% 5 to <10 |jg/dL;
5.1% >10 |jg/dL
Dental caries
Number of teeth with at
least one decayed or
filled surface as detected
by trained examiners
Age at outcome:
2 to 6 yr
Sex, race/ethnicity,
age, urban status,
census region,
poverty index, family
education, ETS
exposure, birth
weight, breastfed,
dental visit, and
parental perception
of oral health
PR for Decayed and
Filled Surfaces
<2 |jg/dL:
Ref.
2-5 |jg/dL:
1.84 (1.36, 2.50)
5-10 |jg/dL:
2.14 (1.36, 3.36)
>10 |jg/dL:
1.91 (1.17, 3.11)
BLL = blood lead level; BMD = bone mineral density; BMI = body mass index; CHEER = Children's Health and Environmental Research; CI = confidence interval; C2C = serum
cleavage neoepitope of type II collagen; COMP = cartilage oligomeric matrix protein; CPU = carboxypropeptide of type II collagen; DMFS = delayed, missing, and filled surfaces;
DMFT = decayed, missing, and filled teeth; DXA = Dual-energy X-ray absorptiometry; ELEMENT = Early Life Exposure in Mexico to Environmental Toxicants; ETS = environmental
tobaccos smoke; GFAAS = Graphite furnace atomic absorption spectrometry; ICP-MS = inductively coupled plasma mass spectrometry; KARE = Korean Association Resource;
KNHANES = Korea National Health and Nutrition Examination Survey; NHANES = National Health and Nutrition Examination Survey; NR = not reported; OA = osteoarthritis;
OR = odds ratio; Pb = lead; PR = prevalence ratio; rOA = radiographic osteoarthritis; sxOA = symptomatic osteoarthritis; SD = standard deviation; Q = quartile; yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
fStudies published since the 2013 Integrated Science Assessment for Lead.
9-126
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Table 9-12 Animal toxicological studies of exposure to Pb and musculoskeletal effects
Study
Species Timing of Exposure
(Stock/Strain), n, Sex Exposure Details
BLL as Reported
(Hg/dL)
Endpoints Examined
Beieretal. (2017)
Mouse (C57BL.6),
0 ppm Pb, M/F, n = NR
100 ppm Pb, M/F,
n = NR
PND 240 Oral, 0.17 ± 0.19 ng/dLfor Serum Protein Levels of Dickkopf-1, Serum Protein Levels of
drinking 0 ppm, Sclerostin (scl), Serum Protein Levels of C-terminal telopeptide
water 58.67 ± 4.61 ng/dL (CTx-1), Serum Protein Levels of type 1 procollagen (P1NP),
for 100 ppm - Energy to Femur Failure (Males, 8 moo), Femur Yield Load /
PND 240 Maximum Load (Males, 8 moo), Maximum Femur Stiffness
(Males, 8 moo), Osteoclast Surface/Bone Surface (Oc.S/BS) by
Micro-Computed Tomography (microCT), Osteoclast
Number/Trabecular Area (N.Oc/Tb.Ar) by Micro-Computed
Tomography (microCT), Osteoblast Number/Trabecular Area
(N.Ob/Tb.Ar) by Micro-Computed Tomography (microCT),
Adipocyte size (Ad Size) by Micro-Computed Tomography
(microCT), Adipocyte Volume/Total Volume (AV/TV) by Micro-
Computed Tomography (microCT), Bone Volume to Total Volume
(BV/TV) by Micro-Computed Tomography (microCT)
Beieretal. (2016)
Mouse (C57BL.6), PND 30, Oral,
0 ppm Pb, F, 200 ppm 90,180, drinking
Pb, F, 500 ppm Pb,/F 360 water
0 ng/mL for 0 ppm, Femur Length, Areal Bone Mineral Density (aBMD), Bone Mass,
50 ng/mL for Bone Weight, Body Fat, Femur Diameter, P1NP (ng/mL),
100 ppm, 100 ng/mL TRAP5b (U/L), CTx (ng/mL), Calcitonin (pg/mL), 17 beta-estradiol
for 300 ppm - (ng/mL), Dkk-1 (ng/mL), Femoral BV/TV, Tb.N, Tb.Sp, Conn.D,
PND 28 SMI, Cort Th, Cort BA, Tb Extension, Bone Strength, Beta-
Catenin Protein Levels, TNF-Alpha Protein Levels, NF-kB Protein
Levels, b-catenin RT-PCR, Peroxisome Proliferator-Activated
Receptor-c RT-PCR, CD47 RT-PCR, Nuclear Factor Of Activated
T Cells RT-PCR, CTSK RT-PCR
aBMD = areal bone mineral density; AV/TV = adipocyte volume/total volume; BV/TV = bone volume to total volume; CTx-1 = C-terminal telopeptide; mo = month(s);
microCT = Micro-Computed Tomography; NF-kB = nuclear factor kappa B; N.Oc/Tb.Ar = Osteoclast Number/Trabecular Area; NR = not reported; Oc.S/BS = Osteoclast
Surface/Bone Surface; P1 NP = type 1 procollagen; PND = postnatal day; RT-PCR = reverse transcription-polymerase chain reaction; scl = sclerostin; TNF = tumor necrosis factor.
9-127
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Table 9-13
Epidemiologic studies of exposure to Pb and ocular effects
Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Glaucoma
tWanq et al. (2018b)
Veterans Affairs NAS
Bone
Glaucoma
Age, BMI,
HRs for Glaucoma
n = 702
education, job
Incidence
United States
Tibia and patella lead
Incident cases of primary
type, pack-yr,
1991-1999 (Follow-up
Healthy male Veterans at
measured using K-XRF
open-angle glaucoma
diabetes mellitus,
Tibia Pb
through 2014)
time of enrollment in the
Age at measurement:
Mean age: 66.8
identified using validated
systemic
1.65)
Cohort
NAS (1963) and without
criteria to assess medical
hypertension, and
1.28 (0.99,
glaucoma at baseline
records
ocular
(time of bone lead
Mean -
hypertension.
Patella Pb
measurement)
Tibia: 21.7 |jg/g
Patella: 31.0 |jg/g
1.42 (1.11,
1.82)
tPark and Choi (2016)
KNHANES
Blood
Intraocular pressure
Age, sex, smoking
Change in intraocular
n = 8371
status, alcohol
pressure (mmHg):
South Korea
Blood Pb was measured in
Intraocular pressure measured
consumption, job
0.088 (0.06, 0.117)
2008-2012
General population,
venous whole blood using
using a Goldmann applanation
status, education,
Cross-sectional
>20 yr old with no history
GFAAS
tonometer
residence,
of glaucoma
Age at measurement:
hypertension
>20 yr old
Age at outcome:
medication use,
>20 yr old
and family history
GM: 2.19 |jg/dL
of glaucoma
+Lin et al. (2015)
KNHANES
Blood
Glaucoma
Age, sex, exercise,
OR for Glaucoma
n = 2680
and ferretin and
Prevalence13:
South Korea
Blood Pb was measured in
Presence of glaucoma was
aspartate
1.04 (0.84, 1.29)
2008-2009
General population,
venous whole blood using
assessed by testing of visual
aminotransferase
Cross-sectional
>19 yr old with no history
GFAAS
function using frequency-
levels
of retinal disease or
Age at measurement:
doubling technology.
stroke
>19 yr old
Age at outcome:
Mean -
19 yr old
w/ glaucoma: 2.70 |jg/dL
w/o glaucoma: 2.52 |jg/dL
9-128
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
+Lee etal. (2016)
KNHANES
n = 5198
Blood
Glaucoma
Age group, region
of residence,
ORs for Glaucoma
Prevalence13
South Korea
Blood Pb was measured in
Presence of glaucoma was
occupation,
2008-2012
General population,
venous whole blood using
assessed by testing of visual
education level,
Normal IOP
Cross-sectional
>19 yr old without a
GFAAS
function using frequency-
smoking status,
history of glaucoma or
Age at measurement:
doubling technology.
hypertension,
0.93 (0.65, 1.34)
age-related macular
>19 yr old
family history of
degeneration
Age at outcome:
glaucoma, and
Low-Teen IOP
GM -
>19 yr old
IOP
1.16 (0.74, 1.83)
No Glaucoma: 2.32 |jg/dL;
Glaucoma: 2.28 |jg/dL
High-Teen IOP
0.65 (0.36, 1.18)
Age-Related Macular Degeneration
tPark etal. (2015)
South Korea
2008-2011
Cross-sectional
KNHANES
n = 3865
General population,
>40 yr old
Blood
Blood Pb was measured in
venous whole blood using
GFAAS
Age at measurement:
>40 yr old
Mean: 2.69 |jg/dL
Age-related macular
degeneration
Macular degeneration was
assessed using retinal
photographs. Photographs
were graded at least twice
using a standardized protocol.
Age at outcome:
>40 yr old
Age, sex, smoking Early-Stage AMD (OR):
status, occupation,
residence,
household income,
anemia, BMI
1.12 (1.02, 1.23)
Late-Stage AMD (OR):
1.25 (1.05, 1.50)
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tHwana et al. (2015)
South Korea
2008-2012
Cross-sectional
KNHANES
n = 4933
General population,
>40 yr old
Blood
Blood Pb was measured in
venous whole blood using
GFAAS
Age at measurement:
>40 yr old
Mean: 3.15 |jg/dL
Quintile
1
<1.75 |jg/dL
Quintile
2
1.75-2.25 |jg/dL
Quintile
3
2.25-2.73 |jg/dL
Quintile
4
2.73-3.38
Quintile
5
>3.38 |jg/dL
Age-related macular
degeneration
Macular degeneration was
assessed using retinal
photographs. Photographs
were graded twice using a
standardized protocol.
Age at outcome:
>40 yr old
NA
ORs (Early-Stage
AMD; Quintiles)
Q1
Q2
Q3
Q4
Q5
Reference
1.04 (0.62, 1.73)
1.14 (0.70, 1.84)
1.26 (0.78, 2.06)
1.55 (0.94, 2.53)
Men Only:
Q1
Q2
Q3
Q4
Q5
Reference
0.66 (0.31, 1.40)
1.32 (0.68, 2.56)
0.80 (0.40, 1.60)
1.32 (0.68, 2.54)
Women Only:
Q1: Reference
Q2: 1.72 (0.86, 3.46)
Q3: 1.83 (0.90, 3.73)
Q4: 1.41 (0.72, 2.77)
Q5: 1.92 (1.06, 3.48)
9-130
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tWu et al. (2014)
United States
2005-2008
Cross-sectional
NHANES
n = 5390
General population,
>40 yr old
Blood
Blood Pb was measured in
venous whole blood using ICP-
MS
Age at measurement:
>40 yr old
GM: 1.61 |jg/dL; Median:
1.77 |jg/dL
75th: 2.61 pg/dL
Max: 26.8 pg/dL
Age-related macular
degeneration
Macular degeneration was
assessed using retinal
photographs. Photographs
were graded twice using a
standardized protocol.
Age at outcome:
>40 yr old
Age, aged-
squared, gender,
race, education,
BMI, pack-yr
ORs for AMD
Prevalence (Quartiles)
Q1
Q2
Q3
Q4
Reference
0.86 (0.60, 1.22)
1.00 (0.68, 1.48)
0.86 (0.59, 1.26)
Quartile 1
Quartile 2
Quartile 3
Quartile 4
0.18-1.2 pg/dL
1.21-1.77 pg/dL
1.78-2.61 pg/dL
2.62-26.8 pg/dL
Other Ocular Effects
(Schaumberq et al..
2004)
United States
1991-1999 (Follow-up
through 2002)
Cohort
Veterans Affairs NAS
n =642
Healthy male Veterans at
time of enrollment in the
NAS
Bone
Tibia and patella lead
measured using K-XRF
Age at measurement:
Mean age: 69 yr
Median -
Tibia: 20 pg/g
Patella: 29 pg/g
Cataract
Documentation for either eye
of cataract surgery or a
cataract, graded clinically as
3+ or higher on a 4-point
scale, diagnosed either after
or within 1 yr prior to bone
lead measurement
Age at outcome:
>60 yr old
Age, pack-yr of
cigarette smoking,
BLLs, diabetes,
and dietary intake
of vitamin C,
vitamin E, and
carotenoids
OR for Cataract
Highest exposure
quintile v lowest
Tibia: 3.19 (1.48, 6.90)
Patella: 1.88 (0.88, 4.02)
9-131
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Reference and Study
Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tWana etal. (2016)
United States
1999-2008
Cross-sectional
NHANES
n = 9763
General population,
>50 yr old
Blood
Blood Pb was measured in
venous whole blood using AAS
(1999-2002) and GFAAS
(2003-2008)
Age at measurement:
50+ yr old
Cataract surgery
Self-reported cataract surgery
Age at outcome:
>50 yr old
Age, race, gender, OR for Cataract
education,
diabetes mellitus,
BMI, serum
cotinine, and pack-
yr
Surgery per doubling
of BLL:
0.97 (0.88, 1.06)
GM: 1.97 |jg/dL
tJunq and Lee (2019)
South Korea
2010-2012
Cross-sectional
KNHANES
n = 23376
General population,
>40 yr old
Blood
Blood Pb was measured in
venous whole blood using
GFAAS
Age at measurement:
>40 yr old
GM -
Male: 2.82 pg/dL;
Female: 2.05 pg/dL
Dry eye disease
Self-reported symptoms of dry
eye disease
Age at outcome:
>40 yr old
Age, sex, smoking ORs for Dry Eye
status, alcohol
consumption,
region, education,
occupation, family
income, family
history of
ophthalmologic
disease, and
history of
ophthalmologic
surgery
Disease Prevalence
(Tertiles)
T1: Reference
T2: 1.12 (0.85,
T3: 0.79 (0.56,
1.48)
1.1)
T1
T2
T3
<2.03 pg/dL
2.03-2.82 pg/dL
>2.82 pg/dL
AAS = atomic absorption spectrometry; AMD = age-related macular degeneration; BLL = blood lead level; BMI = body mass index; GFAAS = Graphite furnace atomic absorption
spectrometry; GM = geometric mean; HR = hazard ratio; ICP-MS = inductively coupled plasma mass spectrometry; IOP = intraocular pressure; KNHANES = Korea National Health
and Nutrition Examination Survey; K-XRF = K-shell X-ray fluorescence; NA = not available; NAS = Normative Aging Study; NHANES = National Health and Nutrition Examination
Survey; OR = odds ratio; Pb = lead; Q = quartile; T# = fertile #; yr = year(s).
aEffect estimates are standardized to a 1 pg/dL increase in BLL or a 10 pg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bPer natural log unit increase in pg/dL of blood Pb.
fStudies published since the 2013 Integrated Science Assessment for Pb.
9-132
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Table 9-14 Animal toxicological studies of Pb exposure and ocular effects
Study Species (Stock/Strain), n, Timing of Exposure Exposure Details BLL as Reported (pg/dL) Endpoints Examined
Shen et al. (2016) Rat (Sprague Dawley), 0 ppm
Pb, M, n = 12 (BLL), n = 6
(other endpoints)
55 ppm Pb (0.01%), M, n = 12
(BLL), n = 6 (other endpoints)
109 ppm Pb (0.02%), M,
n = 12 (BLL), n = 6 (other
endpoints)
Blot Protein Levels of
Occludin, Western Blot Protein
Levels of Claudin-5, Western
Blot Protein Levels of pAkt
(Ser473), Western Blot Protein
Levels of pAkt (Thr308)
BLL weeks 1, 2, 3, 4, Oral, drinking water
5, 6; Other
Endpoints week 6
1.11 ±0.08 |jg/dL for 0.00%
12.58 ±2.42 |jg/dL for 0.01 %
19.00 ±2.59 |jg/dL for 0.02%
Retinal Thickness, Blood-
Retina-Barrier Permeability,
Occludin Protein Levels,
Claudin 5 Protein Levels,
Immunofluorescence Protein
Levels of Occludin,
Immunofluorescence Protein
Levels of Claudin 5, Western
Perkins et al. Mouse (C57BL.6), Bcl-xL
(2012) Transgenic ((C57BL.6),
Background)Wild Type 0.0%
Pb Acetate, M/F, n = 3 to 7,
varying between groups and
between assays
Wild Type 0.015% Pb Acetate,
M/F, n = 3 to 7, varying
between groups and between
assays
Transgenic 0.0% Pb Acetate,
M/F, n = 3 to 7, varying
between groups and between
assays
Transgenic 0.015% Pb
Acetate, M/F, n = 3 to 7,
varying between groups and
between assays
BLL PND 21,
PND60
Other Endpoints
PND 60 to 70
Oral, drinking water
1.9 ± 1.0 pg/dl for 0.0%,
20.6 ± 4.7 |jg/l for 0.015% Pb-
PND21
3.6 ± 1.8 pg/dl for 0.0%,
5.6 ± 2.7 |jg/l for 0.015% Pb-
PND 60
Conventional Transmission
Electron Microscopy (TEM) of
Cell and Organelle Structure,
Three-Dimensional Electron
Microscope Tomography of
Cell and Organelle Structure,
Mitochondrial Cristae
Measurements in Rod
Spherules, Mitochondrial
Cristae Measurements in Cone
Pedicles, Mitochondrial Crista
Junction Diameter and Density
in Rod Spherules,
Mitochondrial Crista Junction
Diameter and Density in Cone
Pedicles, Photoreceptor and
Synaptic Terminal Oxygen
Consumption (Light-Adapted)
BLL = blood lead level; CI = confidence interval; F = female; M = male; pAkt = phosphorylated Akt; Pb = lead; PND = postnatal day; TEM = transmission electron microscopy.
9-133
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Table 9-15
Epidemiologic studies of Pb exposure and respiratory effects
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
Children and Adolescents
tMadriqal et al. (2018)
United States
2011-2012
Cross-sectional
NHANES
n:1234
Children and adolescents
aged 6-17 yr
Blood Pb measured in
venous whole blood
using ICP-MS.
Age at measurement:
6-17 yr old
Median:
0.56
pg/dL
25th
percentile:
0.44
pg/dL
75th
percentile:
0.85
pg/dL
Pulmonary function:
FEVi, FVC, FEVi:
FVC, and FEF25% -75%
Spirometry was
performed in the
standing position
using a standardized
protocol according to
the recommendations
of the American
Thoracic Society for
FEV1 and FVC.
Age at outcome:
6-17 yr old
Age, sex, race, height,
family poverty-income
ratio, serum cotinine, use
of anti-asthmatic,
bronchodilator, or inhaler
medications
Change in lung function
parameters across
blood Pb quartiles
FEVi
Q1: Ref.
Q2: 4.8 (-98.3, 107.8)
Q3: 22.3 (-49.3, 93.9)
Q4: 41.9 (-46.9, 130.6)
FVC
Q1: Ref.
Q2: 1.6 (-88.5, 91.7)
Q3: 23.8 (-46.4, 94.0)
Q4: 45.5 (-49.2, 140.2)
FEVr.FVC
Q1: Ref.
Q2: 0.0003 (-0.01, 0.01)
Q3: -0.001 (-0.01, 0.01)
Q4: 0.002 (-0.01, 0.02)
FEF25%-75%
Q1: Ref.
Q2: -8.1 (-229.8, 213.7)
Q3: -28.9 (-160.5,
102.7)
Q4: 0.71 (-193.1, 192.5)
9-134
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tZenq et al. (2017)
Guiyu, Xiashan, and
Haojiang
Guangdong Province,
China
November -
December 2013
Cross-sectional
Preschool children aged 5-
7 yr
n = 206 (n = 100 from Guiyu,
n = 54 from Xiashan, n = 52
from Haojiang)
Blood Pb measured in
venous whole blood
using GFAAS
Age at measurement:
5-7 yr old
Median
Exposed (Guiyu):
5.53 |jg/dL
Unexposed (Xiashan
and Haojiang):
3.57 |jg/dL
75th Percentile:
Exposed: 7.04 |jg/dL
Unexposed: 4.86 |jg/dL
Lung function
parameters: FVC and
FEV1
Spirometry was
conducted with a
portable spirometer;
results of three
readings were
recorded and the
highest FVC and
FEV1 was used in the
analysis
Age at outcome:
5-7 yr old
Age, gender, height,
family member daily
smoking, family income
level, parental education
level, daily outdoor play
time, and living area
Change in lung function
parameters per In-unit
increase in blood Pb
(Hg/dL)
FEVi (mL)
-15 (-93, 63)
FVC (mL)
-29 (-100, 43)
tLittle et al. (2017)
Legnica-Glogow
District
Poland
1995 and 2007
Cross-sectional
Polish schoolchildren aged
10-15 yr
n = 184 male
n = 189 female
Blood Pb measured in
venous whole blood
using GFAAS
Age at measurement:
10-15 yr
FVC
A Spiro ProVR unit
was used to measure
pulmonary function.
FVC was computed by
the instrument as a
percentage of gender-
, age- and height-
specific normative
data.
Adjusted for height
Change in FVC (mL) per
logio-unit increase in
blood Pb (|jg/dL)
Boys
-5.1 (-13.9, 3.7)
Girls
-12.9 (-23.2, -2.6)
Age at outcome:
10-15 yr
9-135
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tZenq et al. (2016) Children age 3-8
Guiyu and Haojiang
China
December 2012 to
January 2013
Cross-sectional
n = 470 children
n = 170 from Haojiang and
n = 300 from Guiyu)
Blood Pb measured in
venous whole blood
using GFAAS.
Age at measurement:
3-8 yr old
Medians
Guiyu: 6.24 |jg/dL
Haojiang: 4.75 |jg/dL
75th: BLL:
Guiyu: 8 |jg/dL
Haojiang: 5.76 |jg/dL
Respiratory
symptoms: wheeze,
cough, dyspnea, and
phlegm
The respiratory
symptoms such as
wheeze, cough,
phlegm, and dyspnea
were defined by the
standard
questionnaire from the
European Community
Respiratory Health
Survey (ECRHS)
Age at outcome:
3-8 yr old
Age, gender, passive
smoking, living in Guiyu,
whether use home as
workshop, whether home
close to e-waste recycling
site, and whether child
contact e-waste
OR (>5 [jg/dL vs.
<5 [jg/dL blood Pb)
Wheeze
0.64 (0.32, 1.27)
Dyspnea
0.64 (0.23, 1.79)
Cough
0.95 (0.6, 1.52)
Phlegm
1.2 (0.72, 2.01)
Adults
tPaketal. (2012)
Shiwha and Banwol
Korea
2005 and 2007
(Shiwha) and 2006
and 2008 (Banwol)
Cohort
Shiwha and Banwol
Environmental Health Cohort
(SBEHC)
Men and women over the age
of 30 residing in Shiwha or
Banwoi and completed both
pulmonary function tests
during cycle 1 (2005-2006)
and cycle 2 (2007-2008)
Blood Pb measured in
venous whole blood
using GFAAS
GM (GSD):
Cycle 1: 1.55 (1.76)
pg/dL
Cycle 2: 1.96 (1.66)
pg/dL
FEVi and FVC
Pulmonary function
was measure via
spirometry
Age at outcome: 30+
Age, sex, baseline height,
baseline FVC,
methacholine, cotinine
level
Accelerated FVC
Decline
177.0 (24.1, 329.9)
Accelerated FEVi
Decline
107.0 (-0.8, 214.8)
n = 263 (n = 112 males)
9-136
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Reference and
Study Design
Study Population Exposure Assessment Outcome
Confounders
Effect Estimates and
95% Clsa
tLeem et al. (2015)
Korea
2008-2012
Cross-sectional
KNHANES
n = 5972
Adults >20 yr who completed
spirometry and had blood
measurements
Blood Pb measured in
venous whole blood
using GFAAS
Age at measurement:
20+
Mean BLL
non-OLF: 2.36 |jg/dL
OLF: 2.77 pg/dL
Obstructive lung
function (OLF)
Spirometry was used
for lung function. OLF
was defined as
FEV1/FVC <0.7
Age at outcome:
20+
Age, sex, BMI, and
smoking status
Change in lung function
parameters per In-unit
increase in blood Pb
(Hg/dL)
FEVi (mL)
0 (-116, 116)
FVC (mL)
9 (-3, 21)
FEVi/FVC (%)
-0.002 (-0.004, 0)
9-137
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and
95% Clsa
tRokadia and
Aaarwal (2013)
United States
2007-2010
Cross-sectional
NHANES
n = 9575 (1164 OLD and 8411
non-OLD)
Serum Pb measured
from venous whole blood
samples using ICP-MS
Age at measurement:
General population; >18 yr old <| 8-79 yr
Geom. mean (SE)
non-OLD: 1.18 (1.0)
|jg/dL OLD:
pg/dL
1.73 (1.02)
Obstructive lung
disease (OLD)
Spirometric data were
collected from
NHANES participants;
Participants with OLD
were defined as FEV
1 /FVC <0.7; Mild
OLD: FEV1 = 80%
predicted; Moderate-
severe OLD:
FEV1 < 80% predicted
Age at outcome:
18-79 yr
Age, sex, race, BMI,
chronic kidney disease,
diabetes, hyperlipidemia,
hypertension, stroke,
coronary artery disease,
smoking, serum C-
reactive protein
concentration, and serum
cotinine concentration
ORs for OLD
Prevalence
All OLD
1.94 (1.10,
Mild OLD
1.21 (0.55,
3.42)
2.66)
Moderate-Severe OLD
3.49 (1.70, 7.16)
BLL = blood lead level; BMI = body mass index; CI = confidence interval; ECRHS = European Community Respiratory Health Survey; FEF = forced expiratory flow; FEV1 = forced
expiratory volume; FVC = forced vital capacity; GFAAS = graphite furnace atomic absorption spectrometry; GM = geometric mean; GSD = gestational sac diameter; ICP-
MS = inductively coupled plasma mass spectrometry; KNHANES = Korea National Health and Nutrition Examination Survey; NHANES = National Health and Nutrition Examination
Survey; OLD = obstructive lung disease; OLF = obstructive lung function; OR = odds ratio; Pb = lead; SBEHC = Shiwha and Banwol Environmental Health Cohort; Q = quartile;
yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
fStudies published since the 2013 Integrated Science Assessment for Lead.
9-138
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Table 9-16 Animal toxicological studies of exposure to Pb and respiratory effects
Study Species (Stock/Strain), n, Sex Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (pg/dL) b
Endpoints
Examined
Dumkova et al. (2017) Mouse (ICR) NR
experiment 1
Control (clean air), F, n = 5
1.23 x 10s PbO particles/cm3, F, n = 5
experiment 2
Control (clean air), F, n = 5
0.956 x 10s PbO particles/cm3, F, n = 5
Mice were exposed to
PbO NPs 24 hr/d for
6 wk.
<11 ng/g for control
(<1.166 pg/dL)
132 ng/g for Pb-exposed
(13.992 pg/dL)
IHC, Histology
Dumkova et al. Mouse (CD1) (ICR) NR
(2020b)
Control (clean air), F, n = 10 (2 wk, 6 wk,
11 wk)
PbO, F, n = 10 (2 wk, 6 wk, 11 wk)
PbO recovery, F, n = 10 (6 wk PbO, 5 wk
clean air)
174 ng/g PbO 11 wk
(17.4 |jg/dL)
27 ng/g PbO recovery
(6 wk/clean air 5 wk)
(2.7 pg/dL)
Mice were exposed to <3 ng/g in control (2 wk, 6 wk, Western blot,
clean air or PbO np 11 wk) (<0.3 pg/dL) Histology, IHC,
24 hr/d 7 d/wkfor2 wk, PCR
6 wk, or 11 wk. a
recovery group was
exposed to PbO for 6 wk
and then clean air for
5 wk (11 wk total) 148 ng/g PbO 6 wk
104 ng/g PbO 2 wk
(10.4 pg/dL)
(14.8 pg/dL)
9-139
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Study
Species (Stock/Strain), n, Sex
Timing of
Exposure
Exposure Details
(Concentration,
Duration)
BLL as Reported (pg/dL) b
Endpoints
Examined
Dumkova et al.
(2020a)
Mouse (ICR)
Control (clean air), F, n = 10 (d 3, 2 wk,
6 wk, 11 wk)
Pb(N03)2 (68.6 pg/m3), F, n = 10 (d 3,
2 wk, 6 wk, 11 wk)
Recovery (Pb(N03)2 68.6 |jg/m3), F,
n = 10 (6 wk Pb/5 wk recovery)
6 wk - 8 wk Mice were exposed to
at start Pb(N03)2 np or clean air
24 hr/d, 7 d/wk for 3 d,
2 wk, 6 wk, or 11 wk. To
assess recovery, a
separate group of mice
were exposed to
Pb(N03)2 for 6 wk and
then clean air for 5 wk.
<3 ng/g for control at all
timepoints (d 3, 2 wk, 6 wk,
11 wk) (<0.3 pg/dL)
31 ng/g for Pb(N03)2 d 3
(3.1 pg/dL)
40 ng/g for Pb(N03)2 2 wk
(4.0 pg/dL)
47 ng/g for Pb(N03)2 6 wk
(4.7 pg/dL)
PCR, Histology,
IHC
85 ng/g forPb(N03)2 11 wk
(8.5 pg/dL)
10 ng/g forPb(N03)2
exposure 6 wk and clean air
for 5 wk (1.0 pg/dL)
BLL = blood lead level; BMI = body mass index; d = day(s); hr = hour(s); IHC = immunohistochemistry; NP = nanoparticle; NR = not reported; Pb = lead; Pb(N03)2 = lead nitrate;
PbO = lead monoxide; PCR = polymerase chain reaction; wk = week(s).
9-140
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Table 9-17
Epidemiologic studies of Pb exposure and total mortality
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Menke et al. (2006)
NHANES III 1988-1994,
mortality follow-up in 2001
-12 yr of follow-up
Cohort
NHANES III
n = 13,946, >20 yr
Average individual
born -1946
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
Mean: 2.58
Tertiles
T1 < 1.93
T2 1.94-3.62
T3 > 3.63
Age of measurement
Mean 44.4
All-cause mortality
Cox proportional hazard
regression analysis adjusted
age, race/ethnicity, sex, urban
residence, cigarette smoking,
alcohol consumption,
education, physical activity,
household income,
menopausal status, BMI,
CRP, TC, diabetes mellitus,
hypertension, GFR category
HR
All-cause 1.09 (1.05, 1.14)
Schober et al. (2006) NHANES III
n = 9,686, >40 yr
NHANES III 1988-1994,
mortality follow-up in 2006 Average individual
-8.55 yr of follow-up
born in or before
-1951
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
Cohort
T1
T2
T3
<5 (median 2.6)
5-9 (median 6.3)
>10 (median 11.8)
All-cause mortality
Cox proportional hazard
regression analysis adjusted
for sex, age, race/ethnicity,
smoking, education level.
Did not evaluate BMI or
cormorbidities
HR
All-cause 1.05 (1.03, 1.08)
Age of measurement
>40 yr
9-141
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Lustberq and Silberqeld
(2002)
NHANES II 1976-1980,
mortality follow-up in 1992
Cohort
NHANES II
n = 4,190, aged 30-
74
Average individual
born -1924
Blood (GFAAS with
Zeeman correction)15
(Hg/dL)
Mean (SD) 14.0 (5.1)
Median: 13
T1: <10
T2: 10-19
T3: 20-29
Age of measurement
Mean (SD) 54.1 (13.2)
All-cause and
circulatory mortality
Cox proportional hazard
regression analysis adjusted
for age, sex, location,
education, race, income,
smoking, BMI, exercise
HR(T1: Referent)0
All-cause
T2: 1.40 (1.16-1.69)
T3: 2.02 (1.62-2.52)
Khaliletal. (2009)
Baltimore, MD and
Monongahela Valley, PA
Blood Pb measured 1990-
1991, mortality follow-up for
-12 yr
Study of
Osteoporotic
Fractures
n = 533
women, ages 65-
87 yr
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
Mean (SD) 5.3 (2.3)
Range 1-21
Age of measurement
Mean 70
All-cause mortality
Cox proportional hazards
regression analysis adjusted
forage, clinic, BMI, education,
smoking, alcohol intake,
estrogen use, hypertension,
total hip BMD, walking for
exercise, and diabetes
HR (>8 [jg/dL vs. <8 [jg/dL
blood Pb)c
All-cause: 1.59 (1.02, 2.49)
tLanphear et al. (2018)
United States
1988-1994 mortality follow-
up in 2011
-19 yr of follow-up (IQR
17.6-21.0 yr)
Cohort
NHANES III
n = 14,289 >20 yr
Average individual
born -1947
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
Geometric Mean 2.71
Geometric SE 1.31
10th percentile 1.0
90th percentile 6.7
Age of measurement
Mean 44.1
All-cause, CVD, and Cox proportional hazards
IHD mortality regression analysis adjusting
forage, sex, household
income, ethnic origin, BMI,
smoking status, alcohol
consumption, physical activity,
concentration of cadmium in
urine, blood pressure, healthy
eating index tertiles, HbA1C,
and serum cholesterol
HR
All-cause: 1.06 (1.03, 1.09)
CVD: 1.10 (1.05, 1.15)
IHD: 1.14 (1.08, 1.20)
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tvan Bemmel et al. (2011)
United States
1988-1994, follow-up
through 2007
-7.8 yr of follow-up for
those with low blood Pb
-7.5 yr of follow-up for
those with high blood Pb
Cohort
NHANES III
n = 3,349
Adult age >40 yr
Average individual
born -1932
Blood (GFAAS with
Zeeman correction)
(Hg/dL)
Median
<5 |jg/dL 2.6
>5 |jg/dL 7.5
Age of measurement
<5 |jg/dL 57
>5 |jg/dL 61
All-cause and CVD Cox proportional hazards
mortality
adjusting forage, education,
sex, smoking status, and
race/ethnicity
HR
All-cause
All: 1.04 (0.98, 1.10)
ALADGG 1.03 (0.98, 1.08)
ALADCG/GG 1.09 (0.93, 1.28)
tDuan et al. (2020)
United States
1999-2014, follow-up
through end of 2015
- 7.1 yr of follow-up
NHANES
n = 18,602
Age >20 yr
Average individual
born -1960
Blood (ICP-MS) (|Jg/dL)d
Median (IQR)
1.49 (0.93, 2.31)
Age of measurement
Mean (SD) 45.9 (0.3)
All-cause mortality
Poisson regression analyses
adjusted for: sex, age,
ethnicity, education, poverty-
income ratio (PIR), cotinine
category, BMI, physical
activity, hypertension, and
diabetes
RR
All-cause: 1.39 (1.28, 1.51)
Cohort
9-143
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Referenc^and Study study Population Exposure Assessment Outcome
Confounders
Effect Estimates and 95%
Clsa
tBvun et al. (2020)
Korea
2007-2015, mortality
follow-up in 2018 (between
3-11 yr of follow-up)
Cohort
KNHANES
n = 7,308
Individuals with a
BLL less than
10 |jg/dL, who were
aged 30 yr and over
at the baseline
examination, and
who were not
diagnosed with
cancer or IHD
Average individual
born in or before
-1981
Blood (GFAAS with
Zeeman background
correction) (pg/dL)
Geometric mean: 2.26
Blood Pb tertiles:
All-cause mortality
T1
T2
T3
<1.91
1.91-2.71
>2.71
Age at measurement:
>30 yr
Cox proportional hazard
models adjusted for age and
sex, household income,
education, occupation,
smoking status, drinking
frequency, BMI, and physical
activity, high-lead-containing
food intake (grains,
vegetables, and seafood)
HRC
T1
T2
T3
Reference
2.02 (1.20,
1.91 (1.13,
3.40)
3.23)
9-144
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Referenc^and Study study Population Exposure Assessment Outcome
Confounders
Effect Estimates and 95%
Clsa
tLin etal. (2011)
Taiwan
Years not reported
Cohort (18 mo of follow-up)
n = 927
Taiwanese adult
patients with end-
stage renal disease
(ERSD) on
hemodialysis for
>6 mo, age >18
Baseline blood Pb
(ETAAS) (pg/dL)
Mean: 11.5
Median: 10.4
T1
T2
T3
<8.51
8.51-12.64
<12.64
All-cause, and
Infection-cause
mortality
Age of measurement
Mean (SD) 55.2 (13.5)
Multivariate Cox model
adjusting forage, previous
cardiovascular diseases
(stroke, Ml, PID, congestive
heart failure (CHF)), education
level, hemodialysis vintage,
using fistula, normalized
protein catabolic rate,
hemoglobin, serum albumin,
creatinine, cardiothoracic
ratio, and logarithmic
transformation of high-
sensitivity C-reactive protein
(CRP)
HR(T1: Referent)0
All-cause
T2 2.69 (0.47, 3.44)
T3 4.70 (1.92, 11.49)
Infection-cause
T2 4.33 (0.35, 6.54)
T3 5.35 (1.38, 20.83)
Hemoglobin-corrected:
All-cause:
T2: 3.52 (0.41, 5.01)
T3: 4.98 (1.86, 13.33)
Infection-cause:
T2: 3.02 (0.23, 2.07)
T3: 4.72 (1.27, 17.54)
9-145
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Referenc^and Study study Population Exposure Assessment Outcome
Confounders
Effect Estimates and 95%
Clsa
tTonelli etal. (2018)
Canada
Cohort (2 yr of follow-up)
n = 1,278
Patients on incident
hemodialysis
>18 yr
Plasma Pb (ICP-MS)
(pg/dL)
Deciles
1 0.06
2 0.19
3 0.28
4 0.35
5 0.44
6 0.55
7 0.68
8 0.83
9 1.08
10 1.74
All-cause mortality
Logistic regression adjusting
for age, sex, race/ethnicity,
unemployment prior to
dialysis, yr dialysis initiated,
dialysis duration, predialysis
care, arteriovenous access,
comorbidities (atrial fibrillation,
Ml, BMI, cancer,
cerebrovascular disease,
CHF, lung disease, diabetes,
dementia, hypertension, liver
disease, peripheral vascular
disease, psychiatric disease,
substance misuse), albumin,
and creatinine.
*AII variables were considered
candidate variables and were
included based on stepwise
regression results
Authors indicate a null
relationship between blood
Pb deciles and all-cause
mortality; quantitative
results not reported
9-146
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Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
tHollinqsworth and Rudik
(2021) United States
Quasi-experimental design
Elderly population
(>65 yr)
Assessed the
change in deaths
(National Vital
Statistics System)
occurring among
this age group
before and after the
phaseout of leaded
gasoline in
professional racing
(NASCAR, ARCA).
County-level blood Pb
measurements in
children
All-cause mortality
Difference-in-difference
approach controlling for SES
at the county level (median
income, unemployment rates,
percent minority population),
TRI Pb emissions data
Decline in age-standardized
mortality rate per 100,000
population
Race counties: 91
Border counties: 38
Compared mortality
rates in race-
counties to
bordering counties
Average individual
born in or before
-1942
9-147
-------
Reference and Study
Design
Study Population Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
t(Weisskopf et al.. 2015)
United States
Cohort
Veterans Affairs
NAS
n = 637
Healthy male
Veterans at time of
enrollment in the
NAS (1963) and
without glaucoma at
baseline (time of
bone lead
measurement)
Bone
Patella lead measured
using K-XRF
Age at measurement:
Mean (SD): 67 yr (7 yr)
Patella Tertiles
All-cause mortality
T1
T2
T3
<20 |jg/g
20-31 |jg/g
>31 pg/g
Age at K-XRF, age at K-XRF
squared, smoking, education,
occupation and salary at NAS
entry, mother's education and
occupation, father's education
and occupation
HR (T1 Referent)
T2:
T3:
1.41 (0.86, 2.30)
1.86 (1.12, 3.09)
ARCA = Automobile Racing Club of America; BLL = blood lead level; BMD = bone mineral density; BMI = body mass index; CHF = congestive heart failure; CI = confidence interval;
CHF = congestive heart failure; CRP = C-reactive protein; CVD = cardiovascular disease; ERSD = end-stage renal disease; ETAAS = electrothermal atomic absorption spectrometry;
GFAAS = graphite furnace atomic absorption spectrometry; GFR = glomerular filtration rate; HR = hazard ratio; ICP-MS = inductively coupled plasma mass spectrometry;
IHD = ischemic heart disease; IQR = interquartile range; KNHANES = Korea National Health and Nutrition Examination Survey; K-XRF = K-shell X-ray fluorescence; Ml = myocardial
infarction; mo = month(s); NAS = Normative Aging Study ; NASCAR = National Association for Stock Car Auto Racing; NHANES = National Health and Nutrition Examination Survey;
Pb = lead; PIR = poverty-income ratio; RR = relative risk; SD = standard deviation; SES = socioeconomic status, T# = fertile #; TC = total cholesterol; TRI = Toxics Release
Inventory; wk = week(s); yr = year(s).
aEffect estimates are standardized to a 1 |jg/dL increase in BLL or a 10 |jg/g increase in bone Pb level, unless otherwise noted. For studies that report results corresponding to a
change in log-transformed Pb biomarkers, effect estimates are assumed to be linear within the 10th to 90th percentile interval of the biomarker and standardized accordingly.
bBlood Pb analysis method unclear, assumed based on data source.
°Unable to be standardized.
dUnits assumed to be |jg/dL (written as |jg/L in the paper).
fStudies published since the 2013 Integrated Science Assessment for Pb.
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EPA/600/R-23/375
APDA Environmental Protection Januaiy 2024
M m Agency www.epa.eov/isap
Integrated Science
Assessment for Lead
Appendix 10: Cancer
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
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CONTENTS
DOCUMENT GUIDE 10-iii
LIST OF TABLES 10-v
LIST OF FIGURES 10-vi
ACRONYMS AND ABBREVIATIONS 10-vii
APPENDIX 10 CANCER 10-1
10.1 Introduction and Summary of the 2013 Pb ISA 10-1
10.2 Scope 10-3
10.3 Mechanistic Pathways and Markers of Carcinogenesis 10-4
10.3.1 Introduction 10-4
10.3.2 Animal Models of Carcinogenicity 10-5
10.3.3 Genotoxicity 10-6
10.3.4 Oxidative Stress 10-7
10.3.5 Cell Viability, Cytotoxicity, Apoptosis 10-8
10.3.6 DNA Damage Repair Enzymes and Gene Expression 10-8
10.3.7 Epigenetic Regulation of Gene Expression 10-9
10.3.8 Gene Expression and Extracellular Matrix 10-10
10.3.9 Inflammation 10-10
10.3.10 Summary of Mechanistic Pathways and Markers of Carcinogenesis 10-11
10.4 Cancer Incidence and Mortality 10-11
10.4.1 Epidemiologic Studies of Overall Cancer Incidence 10-12
10.4.2 Epidemiologic Studies of Overall Cancer Mortality 10-12
10.4.3 Epidemiologic Studies of Lung Cancer 10-13
10.4.4 Epidemiologic Studies of Brain Cancer 10-13
10.4.5 Epidemiologic Studies of Breast Cancer 10-14
10.4.6 Epidemiologic Studies of Other Cancer 10-14
10.4.7 Summary of Cancer Incidence and Mortality 10-16
10.5 Biological Plausibility 10-18
10.6 Summary and Causality Determination 10-22
10.7 Evidence Inventories - Data Tables to Summarize Study Details 10-28
10.8 References 10-41
10-iv
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LIST OF TABLES
Table 10-1 Summary of evidence for a likely to be causal relationship between Pb exposure and
cancer 10-25
Table 10-2 Epidemiologic studies of exposure to Pb and cancer effects 10-28
10-v
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LIST OF FIGURES
Figure 10-1 Potential biological pathways for cancer from exposure to Pb. 10-19
10-vi
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ACRONYMS AND ABBREVIATIONS
ALAD S-aminolevulinic acid dehydratase
AQCD Air Quality Criteria Document
APE-1 human AP endonuc lease
BLL blood lead level
BMI body mass index
BW body weight
Cd cadmium
CGI CpG island
CI confidence interval
CLL chronic lymphatic lymphoma
CLL/SLL chronic lymphocytic leukemia/small
lymphocytic lymphoma
CPS Cancer Prevention Study
CRP C-reactive protein
d day(s)
DLBCL diffuse large B-cell lymphoma
EPIC European Prospective Investigation
into Cancer and Nutrition
FL follicular lymphoma
GFAAS graphite furnace atomic absorption
spectrometry
GFR glomerular filtration rate
HR hazard ratio
ICD International Classification of Diseases
ICP-MS inductively coupled plasma mass
spectrometry
IC50 half maximal inhibitory concentration
IARC International Agency for Research on
Cancer
IL interleukin type
ISA Integrated Science Assessment
KNHANES Korea National Health and Nutrition
Examination Survey
LINE long interspersed nuclear elements
In natural log
MM multiple myeloma
MMP matrix metalloproteinase-
NHANES National Health and Nutrition
Examination Survey
NHL non-Hodgkin lymphoma
NSDHS Northern Sweden Health and Disease
Study
NTP National Toxicology Program
OR odds ratio
Pb lead
PCR polymerase chain reaction
PECOS Population, Exposure, Comparison,
Outcome, and Study Design
PIR poverty-income ratio
PND postnatal day
ppm parts per million
PRMT protein arginine methyltransferase
Q quartile
RR relative risk
ROS reactive oxygen species
SCE sister chromatid exchange
SD standard deviation
Se selenium
TK thymidine kinase type
UC urothelial carcinoma
WHO World Health Organization
Zn zinc
10-vii
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APPENDIX 10 CANCER
Summary of Causality Determinations for Pb Exposure and Cancer
This appendix characterizes the scientific evidence that supports the causality
determination for lead (Pb) exposure and cancer. The types of studies evaluated within
this appendix are consistent with the overall scope of the ISA as detailed in the Process
Appendix (see Section 12.4). In assessing the overall evidence, the strengths and
limitations of individual studies were evaluated based on scientific considerations
detailed in Table 12-5 of the Process Appendix (Section 12.6.1). More details on the
causal framework used to reach these conclusions are included in the Preamble to the
ISA (U.S. EPA, 2015). The evidence presented throughout this appendix supports the
following causality conclusion:
Outcome
Causality Determination
Cancer
Likely to be Causal
The Executive Summary, Integrated Synthesis, and all other appendices of this Pb ISA
can be found at https://assessments.epa.gov/isa/document/&deid=359536.
10.1 Introduction and Summary of the 2013 Pb ISA
This appendix evaluates the toxicological and epidemiologic literature related to the potential
contributions of Pb exposure to cancer effects, including cancer incidence and mortality. The 2013
Integrated Science Assessment for Lead (hereinafter referred to as the 2013 Pb ISA) continued to support
the conclusions of the 2006 Pb Air Quality Criteria Document (AQCD) that Pb is a well-established
animal carcinogen (U.S. EPA, 2013, 2006). In the 2013 Pb ISA (U.S. EPA, 2013), the toxicological
literature provided consistent evidence of the carcinogenic potential of Pb and possible contributing
modes of action, including genotoxic, mutagenic, and epigenetic effects. The development of cancer is a
multistep process that involves the progressive accumulation of mutations, leading to upregulation of
oncogenes and loss of function of tumor suppressor genes resulting in uncontrolled cell growth and
invasion of cancer cells within organ tissue. Based on the toxicological literature reviewed in the 2013 Pb
ISA, Pb appears to have some ability to induce DNA damage. Additionally, Pb has the ability to alter
gene expression through epigenetic mechanisms and interact with proteins, which may be another
potential means by which Pb induces carcinogenicity (U.S. EPA, 2013). Pb may act at a post-translational
stage to alter protein structure of zinc (Zn)-finger proteins, which can in turn alter gene expression, DNA
repair, and other cellular functions. In summary, cancer develops from one or a combination of multiple
mechanisms including modification of DNA via epigenetics or enzyme dysfunction and genetic instability
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or mutation. These modifications then provide cancer cells with a selective growth advantage and thus, Pb
may contribute to epigenetic changes and chromosomal aberrations.
Multiple longitudinal epidemiologic studies reviewed in the 2013 Pb ISA (U.S. EPA. 2013)
examined the associations between cancer incidence and mortality and Pb exposures, estimated with
biological measures and exposure databases. The 2013 Pb ISA (U.S. EPA. 2013) reported mixed results
for cancer mortality studies. While a high-quality National Health and Nutrition Examination Survey
(NHANES) study demonstrated an association between blood Pb and increased risk of cancer mortality,
other studies reported weak or null associations. Overall, the epidemiologic studies reviewed in the 2013
Pb ISA were well-conducted with control for important potential confounders such as age, smoking, and
education. The epidemiologic studies of cancer incidence in the 2013 Pb ISA reported no associations
between various measures of Pb and overall cancer incidence. These studies were limited by their
ecological or cross-sectional study designs and a few studies did not collect biological measurements, nor
did they control for potential confounders. Additionally, consistent evidence from animal toxicological
studies demonstrated that Pb exposures can lead to cancer, genotoxicity, or epigenetic modification.
Carcinogenicity in animal toxicology studies of Pb exposure were reported in the kidneys, testes, brain,
adrenals, prostate, pituitary, and mammary glands, albeit at high doses of Pb. Furthermore, based on the
previous existing bodies of evidence, International Agency for Research on Cancer (IARC) has classified
inorganic Pb compounds as "probably carcinogenic to humans"1 and the National Toxicology Program
(NTP) has listed Pb and Pb compounds as "reasonably anticipated to be human carcinogen/'2 Overall, the
consistent and strong body of evidence from toxicological studies on tumor incidence and potential modes
of action, when considered together with the inconsistent epidemiologic evidence, was judged sufficient
to conclude that there is likely to be a causal relationship between Pb exposure and cancer.
'The International Agency for Research on Cancer (IARC) classifies carcinogens into four groups. The
categorization of "probably carcinogenic to humans" (Group 2A) applies when IARC has made at least two of the
following evaluations, including at least one that involves either exposed to humans or human cells or tissues:
limited evidence of carcinogenicity in humans; sufficient evidence of carcinogenicity in experimental animals; and
strong evidence that the agent exhibits key characteristics of carcinogens. If there is inadequate evidence of
carcinogenicity in humans, there should be strong evidence in human cells or tissues that the agent exhibits key
characteristics of carcinogens. More information on IARC cancer classifications can be found in the IARC
Monographs on the Identification of Carcinogenic Hazards to Humans Preamble (IARC. 2019).
2The National Toxicology Program (NTP) prepares the Report on Carcinogens (RoC) on behalf of the Secretary of
Health and Human Services and follows an established, multi-step process for the review and evaluation of selected
substances. The classification of "Reasonably Anticipated to be Human Carcinogen" is defined as limited evidence
of carcinogenicity from studies in humans, which indicates that causal interpretation is credible, but that alternative
explanations, such as chance, bias, or confounding factors, could not adequately be excluded; or there is sufficient
evidence of carcinogenicity from studies in experimental animals, which indicates there is an increased incidence of
malignant and/or a combination of malignant and benign tumors (1) in multiple species or at multiple tissue sites, or
(2) by multiple routes of exposure, or (3) to an unusual degree with regard to incidence, site, or type of tumor, or age
at onset; or there is less than sufficient evidence of carcinogenicity in humans or laboratory animals; however, the
agent, substance, or mixture belongs to a well-defined, structurally related class of substances whose members are
listed in a previous Report on Carcinogens as either known to be a human carcinogen or reasonably anticipated to be
a human carcinogen, or there is convincing relevant information that the agent acts through mechanisms indicating it
would likely cause cancer in humans (NTP. 2023).
10-2
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10.2 Scope
The scope of this appendix is defined by Population, Exposure, Comparison, Outcome, and Study
Design (PECOS) statements. The PECOS statements define the objectives of the review and establish
study inclusion criteria thereby facilitating identification of the most relevant literature to inform the Pb
ISA.3 In order to identify the most relevant literature, the body of evidence from the 2013 Pb ISA was
considered in the development of the PECOS statements for this Appendix. Specifically, well-established
areas of research; gaps in the literature; and inherent uncertainties in specific populations, exposure
metrics, comparison groups, and study designs identified in the 2013 Pb ISA inform the scope of this
Appendix. The 2013 Pb ISA used different inclusion criteria than the current ISA, and the studies
referenced therein often do not meet the current PECOS criteria (e.g., due to higher or unreported
biomarker levels). Studies that were included in the 2013 Pb ISA, including many that do not meet the
current PECOS criteria, are discussed in this appendix to establish the state of the evidence prior to this
assessment. With the exception of supporting evidence used to demonstrate the biological plausibility of
Pb-associated cancer incidence and mortality, recent studies were only included if they satisfied all
components of the following discipline-specific PECOS statements:
Epidemiologic Studies:
Population: Any human population, including specific populations or lifestages that might be at
increased risk of a health effect.
Exposure: Exposure to Pb4 as indicated by biological measurements of Pb in the body - with a
specific focus on Pb in blood, bone, and teeth; validated environmental indicators of Pb
exposure5; or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus
lower levels of the exposure metric (e.g., per unit or log unit increase in the exposure metric,
or categorical comparisons between different exposure metric quantiles).
Outcome: Cancer incidence and cancer mortality.
3The following types of publications are generally considered to fall outside the scope and are not included in the
ISA: review articles (which typically present summaries or interpretations of existing studies rather than bringing
forward new information in the form of original research or new analyses), Pb poisoning studies or clinical reports
(e.g., involving accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions), and risk or benefits analyses (e.g., that apply
concentration-response functions or effect estimates to exposure estimates for differing cases).
4Recent studies of occupational exposure to Pb were considered insofar as they addressed a topic area that was of
particular relevance to the National Ambient Air Quality Standards (NAAQS) review (e.g., longitudinal studies
designed to examine recent versus historical Pb exposure).
5Studies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean
aerodynamic diameter less than or equal to 10 |im3 (PMio) and particulate matter with a nominal mean aerodynamic
diameter less than or equal to 2.5 |im3 (PM2.5) ambient air samples are only considered for inclusion if they also
include a relevant biomarker of exposure. Given that size distribution data for Pb-PM are fairly limited, it is difficult
to assess the representativeness of these concentrations to population exposure [Section 2.5.3 (U.S. EPA. 2013)1.
Moreover, data illustrating the relationships of Pb-PMio and Pb-PNfc.s with blood Pb levels (BLL) are lacking.
10-3
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Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies,
case-control studies, cross-sectional studies with appropriate timing of exposure for the health
endpoint of interest, randomized trials and quasi-experimental studies examining
interventions to reduce exposures.
Experimental Studies:
Population: Laboratory nonhuman mammalian animal species (e.g., mouse, rat, Guinea pig,
minipig, rabbit, cat, dog) of any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages).
Exposure: Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that
results in a (blood lead level) BLL of 30 (ig/dL or below" 7
Comparators: A concurrent control group exposed to vehicle-only treatment or untreated
control.
Outcome: Cancer and cancer-related outcomes, such as genotoxicity, epigenetic, and mutagenic
effects.
Study design: Controlled exposure studies of animals in vivo. In vitro mechanistic studies are
supplemental evidence.
10.3 Mechanistic Pathways and Markers of Carcinogenesis
10.3.1 Introduction
The 2013 Pb ISA (U.S. EPA. 2013) reported consistent positive evidence from multiple animal
chronic Pb exposure studies ranging in duration between 18 months to 2 years as well as from animal
studies involving windows of Pb exposure such as gestation and lactation leading to cancers in adult
offspring. Additionally, consistent mechanistic and genotoxicity evidence for cellular and DNA damage
from multiple lines of evidence (human and animal in vitro models) provided further support for
mechanistic pathways of Pb inducing carcinogenicity. The mechanistic toxicological literature evaluated
in the 2013 Pb ISA (U.S. EPA. 2013) found that most evidence clearly supports Pb-induced
carcinogenicity in animal models, but the exact chain of events supporting a mode of action has not been
completely characterized. Furthermore, the IARC (IARC. 2006) classified inorganic Pb compounds as
"probably carcinogenic to humans" (Group 2A), while NTP listed Pb and Pb compounds as "reasonably
anticipated to be human carcinogens" (NTP. 2012). The reports from IARC and NTP based their
6Pb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
7This level represents an order of magnitude above the upper end of the distribution of U.S. young children's BLLs.
The 95th percentile of the 2011-2016 National Health and Nutrition Examination Survey (NHANES) distribution of
BLL in children (1-5 years; n = 2,321) is 2.66 (ig/dL (Eganet al.. 2021) and the proportion of individuals with
BLLs that exceed this concentration varies depending on factors including (but not limited to) housing age,
geographic region, and a child's age, sex, and nutritional status.
10-4
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conclusion on evidence primarily from animal cancer bioassays of continuous exposure to Pb. While
no PECOS-relevant animal studies of Pb exposure and cancer have been published since the 2013 Pb
ISA, a number of recent in vitro studies have examined the potential mechanistic pathways by which
Pb exposure could result in cancer initiation and/or promotion. These mechanistic studies are evaluated
in more detail in the sections below: 10.3.2 Animal Models of Carcinogenicity; 10.3.3 Genotoxicity;
10.3.4 Oxidative Stress; 10.3.5 Cell Viability, Cytotoxicity, Apoptosis; 10.3.6 DNA Damage Repair
Enzymes and Gene Expression; 10.3.7 Epigenetic Regulation of Gene Expression; 10.3.8 Gene
Expression and Extracellular Matrix; and 10.3.9 Inflammation.
10.3.2 Animal Models of Carcinogenicity
The toxicological literature reviewed in previous AQCDs established that Pb has been shown to
act as a carcinogen in animal toxicology models, albeit at relatively high concentrations. Chronic oral Pb
acetate exposure for male and female rodents has consistently been shown to be a kidney carcinogen in
multiple separate studies, inducing adenocarcinomas and adenomas after chronic exposure. The kidneys
are the most common target of Pb-induced carcinogenicity (Kasprzak et al.. 1985; Koller et al.. 1985;
Azar et al.. 1973; Van Esch and Kroes. 1969). Other common targets of Pb-induced carcinogenicity
include the testes, brain, adrenals, prostate, pituitary, and mammary gland (IARC. 2006). The typical
cancer bioassays used by IARC or NTP as evidence of Pb-induced carcinogenicity were designed using
rodents, typically males but occasionally both sexes, that were continuously exposed to Pb acetate in
chow (i.e., 1,000 or 10,000 ppm Pb acetate) or drinking water (i.e., 26 or 2,600 ppm Pb acetate) for
18 months to two years in duration, the typical lifespan of a rodent (Kasprzak et al.. 1985; Koller et al..
1985; Azar et al.. 1973; Van Esch and Kroes. 1969).
The 2013 Pb ISA (U.S. EPA. 2013) focused on the importance of exposure windows for Pb-
induced cancer bioassays in animal toxicology models. Gestational and lactational exposure of rats to
inorganic Pb-induced (500, 750 or 1,000 ppm Pb acetate in drinking water) carcinogenicity in adult
offspring (Waalkes et al.. 1995). In another study, Tokar et al. (2010) considered Pb-induced
carcinogenesis in mice with early life Pb exposure (gestation, lactation and continued until 8 weeks of
age) and examined tumorigenesis in homozygous metallothionein I/II knockout mice and their
corresponding wild-type controls (groups of ten mice each). The dams/mothers were exposed by drinking
water to 2,000 or 4,000 ppm Pb acetate in utero, through birth and lactation, and then, postnatally, to
drinking water until 8 weeks old and compared with untreated controls. The Pb-exposed metallothionein
I/II knockout mice had increased testicular teratomas and renal and urinary bladder preneoplasia. The
tumor burdens of Pb-exposed wild-type mice were not statistically significantly different than controls.
The data suggest that metallothionein can protect against Pb-induced tumorigenesis. The study did not
address whether metallothionein in humans would have any impact on Pb-induced carcinogenesis. The
animal toxicology studies show that Pb is a well-established animal carcinogen in studies employing
high-dose Pb exposure over a continuous, extended duration of exposure (i.e., 2 years), which is typical of
10-5
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cancer bioassays. Studies show early-life maternal Pb exposure can contribute to carcinogenicity in
offspring and suggest that metallothionein is protective against cancer in this pathway.
Since the 2013 Pb ISA, there are no new PECOS-relevant animal studies that have examined
cancer endpoints. Several recent in vitro mechanistic studies have examined markers of potential
carcinogenicity pathways as characterized by the IARC 10 key characteristics of carcinogenic
mechanistic pathways (Smith et al.. 2016). These in vitro mechanistic studies, which are categorized as
supplemental under the PECOS criteria and do not abide by the blood Pb cutoff of 30 (ig/dL, are short-
term in nature, and principally inform mechanistic pathways that inform association to Pb exposure (see
Section 10.2). These in vitro mechanistic studies are detailed below in Sections 10.3.3-10.3.9.
10.3.3 Genotoxicity
Multiple toxicological and epidemiologic studies reviewed in the 2013 Pb ISA (U.S. EPA. 2013)
examined the relationship between Pb exposure and DNA and cellular damage. These studies reported
consistent evidence of genotoxicity, oxidative stress, and related gene expressions. Genotoxic effects are
effects from Pb exposure as measured by multiple lines of evidence such as DNA damage repair. In the
case of DNA strand break detection, in vivo and in vitro studies using the comet assay (measured by
multiple indices such as tail length, single cell electrophoresis, and others) yielded multiple positive
results in various species (Yediou et al.. 2010; Nava-Hernandez et al.. 2009; Tapisso et al.. 2009;
Alghazal et al.. 2008; Kermani et al.. 2008; Xu et al.. 2008; Gastaldo et al.. 2007; Xu et al.. 2006). The
toxicological evidence was supported by several epidemiologic studies that reported associations between
blood Pb and DNA and cellular damage (Khan et al.. 2010; Olewinska et al.. 2010; Shaik and Jamil.
2009; Wiwanitkit et al.. 2008; Duvdu et al.. 2005).
Since the 2013 Pb ISA (U.S. EPA. 2013). there have been additional supplemental studies using
comet assays that continue to indicate DNA strand breakage occurs after Pb exposure across multiple
species (Jiang et al.. 2020; Yadav et al.. 2019; Ali. 2018; Nariva et al.. 2018; Siddarth et al.. 2018; Shah et
al.. 2016; Ahmad et al.. 2015; Mckelvev et al.. 2015; Zhang et al.. 2014; Roy et al.. 2013; Shakoori and
Ahmad. 2013). In addition to DNA and cellular damage, there was a recent study of gamma-H2AX foci
formation from the phosphorylation of the Ser-139 residue of the histone variant H2AX, which is an early
cellular response to the induction of DNA double-strand breaks, with Pb exposure increased these foci
formation (Liu et al.. 2018).
The 2013 Pb ISA (U.S. EPA. 2013) noted Pb-induced micronucleus formation in both the
toxicological and epidemiologic studies reviewed (Shaik and Jamil. 2009; Tapisso et al.. 2009; Alghazal
et al.. 2008). The recently published literature contains multiple studies identifying Pb-induced
micronucleus formation in the human lymphoblastoid cell line (Alimba et al.. 2016) and in human
lymphocytes from healthy volunteers (Nariva et al.. 2018; Shah et al.. 2016; Roy et al.. 2013).
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Sister chromatid exchange (SCE), exchanges of homologous DNA material between chromatids
on a chromosome and are a test for mutagenicity or DNA damage as well as other chromosomal
aberrations in toxicological studies, was outlined extensively in the 2013 Pb ISA (U.S. EPA. 2013). In a
study of mice, the SCE in bone marrow was elevated after treatment with Pb acetate and increased in
time, with co-exposure to cadmium (Cd) or Zn further increasing SCE levels (Tapisso et al.. 2009).
Similarly, recent in vitro studies found Pb-induced damage both in cell lines (Alimba et al.. 2016;
Banfalvi. 2014) and in human peripheral blood lymphocytes (Yadav et al.. 2019; Nariva et al.. 2018; Shah
et al.. 2016).
10.3.4 Oxidative Stress
At cellular level, Pb is known to induce oxidative stress either by generation of free radicals or
through depletion of antioxidants (Ercal et al„ 2001). Pb-induced free radicals initiate DNA oxidation and
subsequent DNA damage (Hsu and Guo, 2002) as well as mitochondrial damage and intracellular
depletion of glutathione (Sabath and Robles-Osorio, 2012).
Since the 2013 Pb ISA, multiple studies have investigated Pb-induced oxidative stress and
diverse biomarkers in the context of genotoxicity and carcinogenic mechanisms. All these studies used
in vitro cell culture (human and mammalian animal) systems exposed to either Pb acetate or Pb nitrate of
varied concentrations/doses and durations. Some of these studies also examined the effect of antioxidant
treatment on the reversal of oxidative stress endpoints and of genotoxic endpoints resulting from Pb-
induced oxidative stress. Nariya et al. (2018) observed dose (Pb acetate; 0.379 |ig/m 1 and 37.9 (ig/ml) and
duration (24 or 69 hours) dependent increases in oxidative stress and genotoxicity (chromosomal
aberrations, micronuclei) and reversal of these effects when treated with antioxidant and anti-
inflammatory curcumin (1.43 (ig/ml). Similarly, Yadav et al. (2019) observed reversal of Pb nitrate (50-
350 |ig/m 1 for 24 hours) induced genotoxicity (as assessed by comet assay and sister chromatid exchange)
by pretreatment of human peripheral blood lymphocytes with antioxidant, anti-inflammatory
bioflavonoid, "morin'. at concentrations of 15-60 |ig/m 1.
Three recent studies evaluated Pb-induced oxidative stress and its effects on DNA damage. Liu et
al. (2018) used thymidine kinase (TK) 6 cells exposed to Pb acetate (0-480 mM) for 6-24 hours and
observed the formation of 8-OH guanosine adducts and gamma-H2AX foci, markers of DNA double-
strand breaks. Pottier et al. (2013) also observed a dose dependent (Pb-nitrate; 0-1000 mM) loss of
telomeres in clone B3 of the human EJ30 bladder carcinoma cell line. In these cells, formation of foci
(indicative of cell transformation) was found only above 100 mM Pb. Jiang et al. (2020) observed Pb-
induced DNA damage mediated by oxidative stress and inflammation mechanism in human lung cells at
no-observed-adverse-effect level of 4 (ig/ml Pb. Ali (2018) also observed Pb-induced DNA damage
mediated by oxidative stress in human lung cells at half maximal inhibitory concentration (IC50) dose of
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Pb. Furthermore, the IC50 dose of Pb-induced DNA damage was found to be reversed when treated with
antioxidants (i.e., vitamin E or garlic extract) (Ali. 20 IS).
10.3.5 Cell Viability, Cytotoxicity, Apoptosis
Toxicant-induced oxidative stress, if left uncontrolled or depleted of cellular antioxidant
resources, eventually leads to DNA or chromatin damage and cell death or apoptosis. Since the 2013 Pb
ISA, a limited number of in vitro cell culture studies that observed Pb-induced oxidative stress further
investigated cytotoxicity mechanisms. Ali (2018) found a dose dependent increase in cell viability and
cytotoxicity in association with Pb exposure. In addition, garlic, vitamin E, and the combination mitigated
these effects to different levels. The cytotoxicity was found to be associated with alterations in the
expression of pro-apoptotic genes (bcl2, Bax, P53) and significant increase in Bax/Bcl2 ratio suggesting
their role in an apoptotic mechanism of cytotoxicity. Jiang et al. (2020) also observed a dose-dependent
decrease in cell viability associated with changes in the expression of specific proapoptotic genes (caspase
3, 8, and 9). Similarly, Siddarth et al. (2018) observed increased expression of caspase 3 and an increased
number of annexin V positive cells by flow cytometric analyses, suggesting an apoptotic mechanism for
cell death. These studies also found reversal of these effects when treated with diverse antioxidants (see
Section 10.3.4 on oxidative stress). Ghosh et al. (2018) observed significant Pb chloride (5 mM and
10 mM) induced decreases in cell viability of A549 human lung and MCF-7 human breast cancer cell
lines as assessed by trypan blue exclusion, MTT assay, and neutral red dye uptake methods.
10.3.6 DNA Damage Repair Enzymes and Gene Expression
Cells are equipped with robust, diverse DNA damage response mechanisms consisting of specific
DNA repair pathways to remove damage and effect repair at different stages of the cell cycle. Since the
2013 Pb ISA, two in vitro studies have investigated the role of DNA damage repair enzymes by studying
their expression after Pb exposure. Mckelvey et al. (2015), using the RT2 Profiler polymerase chain
reaction (PCR) array system, found that exposure to Pb nitrate (40 |ig/m 1 and 80 |ig/m 1) impacted diverse
DNA damage and signaling pathways in the HepG2 (human hepatocellular carcinoma) cell line. These
investigations were carried out in the context of protection conferred by diverse chemical forms of
selenium (Se) to Pb-induced DNA damage. The potential role for the changes in genotoxicity was
complemented by the comet assay and other methods (discussed in Section 10.3.1). Both doses of Pb
nitrate led to increased expression of several genes and the study reported differential fold increases
between the 40 |ig/m 1 and 80 |ig/m 1 doses. The two most significant increases were found in the
expression of GADD45G (growth arrest and DNA-damage inducible, gamma) and PPP1R15A (protein
phosphatase 1, regulatory subunit 15 A) by 26- and 12-fold, respectively, in cells exposed at 40 mg/ml.
Smaller increases were reported in cells exposed at 80 mg/ml (4- and 6-fold, respectively). The ATM
gene that functions as a main sensor of DNA damage and is involved in DNA double-strand break (DSB)
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repair was found to be suppressed by Pb nitrate. In this study the protection conferred by diverse Se-based
compounds sodium selenite (Sel-Ni), selenium yeast (SeY), seleno-methionine (Sel-M), and sodium
selenate (Sel-Na) were also investigated in the gene expression of Pb-induced DNA repair enzymes. It
was observed that SeY and Sel-M influenced the Pb-induced expression of LIG1 (ligase I, DNA) and
XRCC3, two important genes involved in the base excision repair pathway, indicating that Pb-induced
oxidative stress might influence the expression and regulation of these enzymes and that these Se
compounds confer protection against it.
10.3.7 Epigenetic Regulation of Gene Expression
The 2013 Pb ISA reported that the ability of Pb to alter gene expression through epigenetic
mechanisms and to interact with proteins may be a means by which Pb induces carcinogenicity (Patch
2013; Li et al.. 2011; Wright et al.. 2010; Pilsner et al.. 2009). Cancer develops from one or a combination
of multiple mechanisms including modification of DNA via epigenetics or enzyme dysfunction and
genetic instability or mutation. These modifications can then provide the cancer cells with a selective
growth advantage, in which Pb may contribute to epigenetic changes and chromosomal aberrations.
Additionally, epigenetic modifications may lead to cancer by altering cellular functions without altering
the DNA sequence. The most studied epigenetic change is methylation alterations. A small number of
studies included in the 2013 Pb ISA show that Pb can induce epigenetic changes, but do not clearly tie
these effects to Pb-induced carcinogenesis and genotoxicity (Patel, 2013; Li et al., 2011; Wright et al.,
2010; Pilsner et al., 2009). Since the 2013 Pb ISA, additional studies have examined Pb-induced
epigenetic modifications and the degree to which these modifications may underlie Pb-induced
carcinogenicity. These studies are discussed below.
The role of epigenetic mechanisms such as DNA methylation (and demethylation), histone
modifications, and non-coding RNAs in the regulation of gene expression is well established. Promoter
methylation of DNA repair genes is a common event in tumorigenesis. Two recent in vitro cell culture
studies investigated the potential effects of Pb on epigenetic regulation of gene expression. Liu et al.
(2018), using methylation-specific PCR (M-PCR) that specifically enhances promoter methylation,
investigated TK-6 cells exposed to Pb acetate at different time points. Expression of several DNA repair
genes (XRCC1, hOGG-1, BRCA1, and XPD) was inhibited in this assay, suggesting a role for alterations
in methylation profiles of these genes.
Histone and non-histone proteins are methylated by a family of protein arginine methyltransferase
(PRMT) enzymes. One of the isoforms of this enzyme, PRMT5, is an oncogene and plays a critical role in
cancer progression by promoting cell proliferation and inhibiting apoptosis; moreover, it is overexpressed
in many forms of human cancers (Dai et al., 2022; Stopa et al., 2015; Bao et al., 2013; Nicholas et al„
2013). Using in vitro culture systems (A549 and MCF-7 cell lines) exposed to Pb chloride (5 and 10 |iM)
for 24 and 48 hours, Ghosh et al. (2018) investigated Pb-induced, global DNA hypomethylation and
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methylation status specific to PRMT5 promoter CpG islands (CGIs). Pb-chloride exposure was found to
reduce global methylation levels and either completely or partially demethylate only the upstream
PRMT5 promoter CGI. Additional confirmational studies using bisulfite sequencing indicated an
approximately five-fold reduction in the methylation by Pb chloride. These two recent studies (Ghosh et
al.. 2018; Liu et al.. 2018) suggest the potential for Pb exposure to alter epigenetic control of gene
expression.
10.3.8 Gene Expression and Extracellular Matrix
A single recent study examined gene expression related to cancer progression as assessed by
epithelial-to-mesenchymal transition and invasiveness in Renca cells, a murine renal cortical
adenocarcinoma cell line (Akin et al.. 2019). In these cells, Pb-induced a concentration-dependent (0,
0.625, 1.25 (j,M) decrease in E- cadherin expression with no alteration in catenin expression, a substantial
increase in matrix metalloproteinase-9 (MMP9; involved in cell migration) expression, significantly
reduced cell aggregates, and increased cell migration and invasion. Pb exposure also enhanced wound
healing in a functional "scratch" assay.
10.3.9 Inflammation
Inflammation is positively associated with the development and progression of cancer (Zhao et
al.. 2021). Two in vitro cell culture studies investigated markers of inflammation after Pb exposure using
a cancer cell line (Jiang et al.. 2020; Lin et al.. 2015). Lin et al. (2015) investigated Pb nitrate-induced
(0.1 |iM) inflammation using human stomach adenocarcinoma cells. Pb nitrate was found to induce
expression of the proinflammatory gene, interleukin type 8 (IL-8), in a time-dependent manner. Detailed
molecular characterization studies on upstream events indicated transcription factor activator protein 1 to
be a major transcription factor responsible for this activation while another transcription factor, NF-kB,
played only a minor role. Lin et al. (2015) conducted additional experiments using promoter reporter
assay. These experiments indicated that induction of IL-8 is mediated by activation of extracellular
regulator kinase 1/2 and epidermal growth factor receptor upstream of extracellular regulated kinase 1/2
pathway, an important mediator of cytokine secretion. The observation of Pb-induced expression of the
proinflammatory cytokines IL-8 and tumor necrosis factor-a in BEAS-2b human lung cells by Jiang et al.
(2020) also confirms the role of inflammation in Pb exposure. Additional experiments suggest that Pb-
induced oxidative stress may be the initial event triggering this response (Yadav et al.. 2019; Nariva et al..
2018).
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10.3.10 Summary of Mechanistic Pathways and Markers of Carcinogenesis
The toxicological literature provides consistent evidence for the carcinogenic potential of Pb, and
the findings of Pb-induced genotoxic, mutagenic, and epigenetic effects are consistent with the
conclusions drawn in the 2013 Pb ISA. Among the toxicological literature reviewed in the 2013 Pb ISA,
laboratory studies in animals consistently report cancer following chronic Pb exposure for 18 months or
two years to high concentrations, such as 10,000 ppm Pb acetate in diet or 2,600 ppm Pb acetate in
drinking water. Chronic Pb exposure to male and female rodents has consistently induced kidney and
brain carcinogenesis in multiple separate studies, inducing various tumors (i.e., adenocarcinomas,
adenomas, and gliomas). Pb has also been shown to cause mammary gland, prostate, adrenal, and
testicular tumors in animals. Developmental Pb acetate exposure also induced tumors in offspring whose
dams received Pb acetate in drinking water during pregnancy and lactation.
In the absence of any new cancer bioassay studies using animal models, much of the toxicological
evidence evaluated here comes from in vitro studies using several mammalian cell culture systems
(micromolar to millimolar concentrations). These studies provide evidence supporting the Pb-induced
activation of diverse mechanistic pathways that are normally associated with carcinogenesis. The new
studies continue to support that exposure to multiple forms of Pb (i.e., Pb ions such as Pb acetate, Pb
nitrate, or Pb chloride) induces cellular oxidative stress that triggers a set of biological pathways leading
to DNA damage, cytotoxicity, and apoptosis. In several cases, the observed effects were exposure related
and were both dose dependent and duration dependent. The molecular alterations are diverse in nature,
including modified expression of various genes, epigenetic regulatory changes, and activation of upstream
mediators for specific oncogenic pathways. Some of the studies also demonstrated that antioxidant
administration prior to (or simultaneous with) treatment with Pb protected against Pb-induced effects.
Studies of DNA damage and repair after Pb exposure, where oxidative stress seems to be involved,
provide additional evidence in support of these observations. In addition, Pb-induced oxidative stress is
implicated in multiple organ (liver and kidney) toxicity in animals and supports a strong role for this
molecular pathway in Pb-induced toxicity and cancer. Most of the biological pathways implicated in Pb
carcinogenesis reviewed here are part of the IARC-identified 10 key characteristics, further supporting
conclusions derived in 2013 Pb ISA (U.S. EPA. 2013).
10.4 Cancer Incidence and Mortality
Recent studies have included epidemiologic evaluations of the associations between Pb exposure
and both specific cancers (such as breast cancer and lymphoid malignancies), and overall cancer (cancer
of any type). Table 10-1 provides an overview of the study characteristics and results for the
epidemiologic studies that reported effect estimates.
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10.4.1
Epidemiologic Studies of Overall Cancer Incidence
The epidemiologic studies reviewed in the 2013 Pb ISA (U.S. EPA. 2013) found no positive
associations between various biological markers of Pb exposure and overall cancer incidence. The few
epidemiologic studies evaluated were limited by the ecologic or cross-sectional study designs.
Additionally, these studies were limited by the lack of biological measurements of Pb and the lack of
adjustment for potential confounders. There were no recent PECOS-relevant epidemiologic studies of
overall cancer incidence and Pb exposure.
10.4.2 Epidemiologic Studies of Overall Cancer Mortality
The 2013 Pb ISA (U.S. EPA. 2013) reviewed several epidemiologic studies that examined the
associations between blood Pb concentrations and cancer mortality. The findings of these studies were
inconsistent. More specifically, the findings were inconsistent among participants from NHANES III. In
one NHANES III analysis, the cohort of 13,946 (n for cancer mortality = 411) was followed for 12 years
and individuals with BLLs greater than 10 (ig/dL were excluded from the study (mean baseline BLL was
2.58 (ig/dL) (Menke et al.. 2006). There were null associations between blood Pb and cancer mortality
(hazard ratio [HR] of highest tertile [>3.63 |ig/dL| compared with lowest tertile [<1.93 |ig/dL|: 1.10
[95% CI: 0.82, 1.47]). Another NHANES III study, which was restricted to individuals 40 years and older
at the time of blood Pb collection and included 9,757 (N for cancer mortality = 543) individuals with all
BLLs (including those greater than 10 (.ig/dL). reported positive associations between blood Pb and
cancer mortality (Schober et al.. 2006). The RRs were 1.69 (95% CI: 1.14, 2.52) for individuals with
BLLs of at least 10 (ig/dL and 1.44 (95% CI: 1.12, 1.86) for BLLs of 5-9 (ig/dL, compared with
individuals with BLLs less than 5 (ig/dL. Overall, while the epidemiologic studies reviewed in the 2013
Pb ISA (U.S. EPA. 2013) were well-conducted longitudinal studies with control for wide range potential
confounders, the studies were limited by the small number of cancer mortality cases, which reduces
precision of the measures of associations.
There are a limited number of recent epidemiologic studies which examined the associations
between exposure to Pb and overall cancer mortality (Table 10-2). Total mortality is discussed in
Section 9.8 in Other Health Effects. Multiple population-based studies found inconsistent associations
between blood Pb concentrations and overall cancer mortality (Bvun et al.. 2020; Duan et al.. 2020; van
Bemmel et al.. 2011). A subset of NHANES III data (1984-1994) that included adults over the age of 40
(n = 3,223), in study participants with elevated BLLs (>5 (.ig/dL). there were null associations with overall
cancer mortality (HR: 1.083 [95% CI: 0.983, 1.194]), compared with those with lower BLLs (<5 (ig/dL)
(van Bemmel et al.. 2011). Furthermore, the hazard ratio was nearly unchanged when the data were
stratified by an S-aminolevulinic acid dehydratase (ALAD) genetic polymorphism (ALADGG) that may
influence a person's susceptibility to lead poisoning. In another NHANES study (1999-2014), which
included adults over the age of 20 (n = 26,056), blood Pb was positively associated with cancer mortality
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(1.47 [95% CI: 1.22, 1.78]) in the fully adjusted models (Duan et al.. 2020). In the 2007-2015 Korea
National Health and Nutrition Examination Survey (KNHANES), Bvun et al. (2020) reported positive
associations between blood Pb and cancer mortality, among the 7,308 study participants, who were at
least 30 years of age at baseline. Compared with the lowest tertile (blood Pb <1.91 (.ig/dL). the HRs for
cancer mortality in the second (blood Pb between 1.91 and 2.71 (ig/dL) and third (blood Pb >2.71 (ig/dL)
tertile of blood Pb were 3.42 (95% CI: 1.65, 7.08) and 2.27 (95% CI: 1.09, 4.70), respectively. The nature
of the concentration response relationship appears to be non-linear, but the imprecision in the estimates
ultimately limits the ability to make any inferences about the relationship.
In summary, there are a limited number of recent epidemiologic studies that examined the
association between blood Pb concentrations and overall cancer mortality (Table 10-2). These recent
studies used exposure data from population-based national surveys linked to mortality records. The
NHANES studies reported null associations between BLLs and overall cancer mortality. The median and
geometric mean of BLLs among the NHANES studies were all below 10 (ig/dL (median range:
1.49 (ig/dL to 7.5 (ig/dL; geometric mean: 2.26 (ig/dL). In the population-based study in South Korea,
there were positive associations with cancer mortality among participants with BLLs less than 10 (ig/dL.
Because the participants in the population-based South Korean study would likely have had higher past
Pb exposures due to when the leaded gasoline was banned in South Korea, uncertainty exists as to the Pb
exposure level, duration, frequency, and timing associated cancer mortality. Additionally, while these
epidemiologic studies were conducted in well-established cohorts, there is uncertainty in their
interpretation because the overall follow-up period was short (<11 years). These studies also had a small
number of cancer mortality cases, which resulted in reduced precision across the studies. There was a lack
of control for some potential influential confounders such as co-morbidities and body mass index (BMI).l
10.4.3 Epidemiologic Studies of Lung Cancer
The epidemiologic studies reviewed in the 2013 Pb ISA of Pb (U.S. EPA. 2013) exposure and
lung cancer reported no evidence of an association. The studies available for review were conducted in
occupational cohorts and only included male study participants, which limits the generalizability of the
results. A few of the studies did not obtain Pb biomarker exposure levels or only used air sampling
measurements. Furthermore, these studies may be confounded by other workplace exposures and
covariates, such as smoking, that were not considered. There were no recent PECOS-relevant
epidemiologic studies of Pb exposure and lung cancer.
10.4.4 Epidemiologic Studies of Brain Cancer
The 2013 Pb ISA (U.S. EPA. 2013) reviewed a few studies of brain cancer and occupational Pb
exposure. The associations between occupational Pb exposure and brain cancer incidence and mortality
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varied depending on the tumor type or genetic variant. The implications of the results from these studies
were limited because they did not have individual-level biological Pb measurements, relied on self-
reported occupational exposure history, and did not control for potential confounding by other workplace
exposures. There were no recent PECOS-relevant epidemiologic studies of Pb exposure and brain cancer.
10.4.5 Epidemiologic Studies of Breast Cancer
The epidemiologic studies reviewed in the 2013 Pb ISA (U.S. EPA, 2013) of Pb exposure and
breast cancer suggested that women with breast cancer may have higher BLLs than those without breast
cancer. These studies were limited by their study designs, small sample sizes, and with one study, the
method of Pb exposure measurement. There were also some inconsistent results among studies that
compared breast tissue Pb concentrations between breast tumor and control samples.
Since the 2013 Pb ISA, a few epidemiologic studies of Pb exposure in blood and breast cancer
have been published (Table 10-2). Gaudet et al. (2019) examined associations of circulating levels of Pb
with breast cancer risk in three case-control studies nested withing three prospective longitudinal cohorts
in the United States, Italy, and Sweden. Among the three cohorts, there were consistent null associations
between circulating BLLs and breast cancer, both when Pb exposure was evaluated continuously
(RR= 1.00) and when categorized into quintiles (RR range: 0.65-1.10). In a cross-sectional study of
NHANES data, Wei and Zhu (2020) reported increased odds of breast cancer across quartiles of BLLs.
The odds of breast cancer were 2.52 (95% CI: 1.35, 4.73) in the second quartile (0.8 - 1.2 (ig/dL), 2.01
(95% CI: 1.05, 3.84) in the third quartile (1.2-1.8 (ig/dL), and 2.63 (95% CI: 1.36, 5.09) in the highest
quartile £1.8 (ig/dL), compared with the lowest quartile (<0.8 (ig/dL).
Overall, the current epidemiologic studies evaluating the associations between breast cancer and
blood Pb reported inconsistent findings, with a cross-sectional NHANES study finding increasing odds of
breast cancer across blood Pb quartiles, while another study using three longitudinal cohorts did not find
associations between breast cancer and blood Pb. The inconsistency in findings may be related to
differences in study design, biomarkers of exposure as Wei and Zhu (2020) measured Pb in whole blood,
while Gaudet et al. (2019) measured Pb levels in stored erythrocytes, timing of exposure (blood draws
were obtained from 1990-2006 in Gaudet et al. (2019), while Wei and Zhu (2020) used data from 2003-
2012), and range of Pb levels.
10.4.6 Epidemiologic Studies of Other Cancer
The epidemiologic literature reviewed in the 2013 Pb ISA (U.S. EPA. 2013) for associations
between Pb exposures and other specific cancers reported varying associations among occupational
cohorts. Positive associations were observed between occupational exposure to Pb and adenocarcinoma of
the esophagus and stomach cancer, but there were inconsistent associations with occupational Pb
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exposure and rectal cancer and occupational leaded gasoline exposure and stomach cancer. These
occupational cohort studies were limited to the study populations consisting of only men, no personal,
biological, or exposure measurements for Pb, and no control for potential confounding by other
occupational exposures. The current epidemiologic literature examining the associations of Pb exposure
and specific cancer outcomes remains limited. Table 10-2 provides an overview of the current
epidemiologic study details.
A single study evaluated the association between BLLs and urothelial carcinoma in a hospital-
based case-control study in China (Chung et al.. 2017). Study participants were recruited between 2011
and August 2014, resulting in 209 cases matched to 417 controls based on age (range: 26-96 years) and
gender. Cases has slightly higher Pb blood levels (mean: 2.81 (ig/dL) than controls (mean: 2.56 (.ig/dL).
There were increased odds of urothelial carcinoma (OR: 1.66 [95% CI: 1.05, 2.61]) in the highest quartile
(>2.99 (ig/dL) of blood Pb compared with the lowest (<1.76 (ig/dL). There was also increased risk of
urothelial carcinoma in the highest tertile of blood Pb (>2.73 (ig/dL) for both current smokers (OR: 1.76
[95% CI: 0.69, 4.46]) and non-smokers (OR: 1.48 [95% CI: 0.91, 2.39]).
In a hospital-based case-control study in China, Lin et al. (2018) examined the BLLs and
associations with gastrointestinal cancers. There were 167 gastrointestinal cancer cases (70 esophageal,
51 gastric, and 46 colorectal), which were newly diagnosed and previously untreated, and 112 controls
included in the study. The BLLs were slightly higher among cases (median: 6.003 (ig/dL) than controls
(median: 5.384 (ig/dL). The 75th percentile of the BLL (9.09 (ig/dL) of cases was used as a cutoff to
assign study participants as either low (<75th percentile) or high (>75th percentile) blood Pb. There was
an increased odds of 2.32 (95% CI: 1.01, 4.94) of gastrointestinal cancers for those with high BLLs,
compared with those with low BLLs. When stratifying by clinical characteristics among cases with high
BLLs (>9.09 (ig/dL, 75th percentile), there were positive, but imprecise associations due to the small
number of cases (i.e., <20 cases) (see Table 10-2).
Kelly et al. (2013) and Deubler et al. (2020) examined the associations between Pb exposure in
blood erythrocytes and lymphoid malignancies, specifically B-cell non-Hodgkin lymphoma (NHL) and
multiple myeloma (MM), in large prospective cohorts in the United States, Italy, and Sweden. Kelly et al.
(2013) conducted a case-control study nested within two prospective cohorts in Italy (n = 84 cases and
n = 84 controls) and Sweden (n = 186 cases and n = 186 controls). Lymphoma cases were identified
between 2-16 years of follow-up and controls were matched on gender, age, center (Italy or Sweden), and
date of blood collection. With increasing quartiles of pre-diagnostic exposure levels of Pb, Kelly et al.
(2013) reported null associations with B-cell NHL (OR: 0.93 [95% CI: 0.43, 2.02]) for the total study
population (both cohorts), and the null associations remained when stratified by sex [OR for males: 0.74
(95% CI: 0.27, 2.04); OR for females: 0.42 (95% CI: 0.12, 1.47)]. When comparing the highest quartile of
pre-diagnostic exposure levels of Pb to the lowest, there was increased odds of 1.63 (95% CI: 0.45, 5.94)
for MM among the total study population, but this association was imprecise due to the small sample size.
There were insufficient numbers to stratify by males, but for females there was no association between
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MM and the highest quartile of pre-diagnostic exposure levels of Pb (OR:0.74 [95% CI: 0.14, 3.83]).
When further stratified by NHL subtype, there were null associations: diffuse large B-cell lymphoma
(OR: 0.60 [95% CI: 0.26, 1.40]), B-cell chronic lymphatic lymphoma (OR:0.71 [95% CI: 0.32, 1.57]),
MM (OR: 1.04 [95% CI: 0.57, 1.90]), and follicular lymphoma (OR: 1.17 [95% CI: 0.52, 2.63]) per one
unit increase in log-transformed pre-diagnostic exposure levels of Pb. There were null associations for
females for MM (OR: 1.28 [95% CI: 0.53, 1.96]), follicular lymphoma (OR: 1.91 [95% CI: 0.54, 6.78]),
diffuse large B-cell lymphoma (OR:0.29 [95% CI: 0.07, 1.18]), or B-cell chronic lymphatic lymphoma
(OR:0.79 [95% CI: 0.17, 3.60]) per one unit increase in log-transformed pre-diagnostic exposure levels of
Pb. There were null associations between males and MM (OR:0.83, 95% CI: 0.35, 1.96), diffuse large B-
cell lymphoma (OR:0.97 [95% CI: 0.35, 2.64]), B-cell chronic lymphatic lymphoma (OR:0.63 [95% CI:
0.23, 1.74]), or follicular lymphoma (OR:0.80 [95% CI: 0.25, 2.55]) per one unit increase in log-
transformed pre-diagnostic exposure levels of Pb.
Deubler et al. (2020) also conducted a case-control study, but among participants of the Cancer
Prevention Study-II Nutritional Cohort (CPS-IINC) to assess the risk of lymphoid malignancies, B-cell
NHL and MM, with pre-diagnostic erythrocyte Pb levels. There were 375 cases and 750 controls. There
were positive associations with overall lymphoid malignancy (RR: 1.088 [95% CI: 1.009, 1.173] per 1-
SD (1.76 (ig/dL) increase of erythrocyte lead concentrations), all B-cell NHL (RR: 1.093 [95% CI: 1.005,
1.19] per 1-SD increase of erythrocyte lead concentrations), and follicular lymphoma (RR: 1.114 [95%
CI: 1.085, 1.798] per 1-SD increase of erythrocyte lead concentrations), but null associations with diffuse
large B-cell lymphoma, chronic lymphocytic leukemia/small lymphocytic lymphoma (CLL/SLL), other
B-cell lymphoma, and MM. When stratified by sex, for males, there were positive associations between
overall lymphoid malignancy (RR: 1.131 [95% CI: 1.027, 1.246] per 1-SD (1.81 (ig/dL) increase in
erythrocyte Pb), all B-cell NHL (RR: 1.151 [95% CI: 1.03, 1.286] per 1-SD increase in erythrocyte Pb),
CLL/SLL (RR: 1.274 [95% CI: 1.016, 1.598] per 1-SD increase in erythrocyte Pb), but null associations
with diffuse large B-cell lymphoma, follicular lymphoma, other B-cell lymphoma, and MM. Among
females, there was a positive association with follicular lymphoma (RR: 2.158 [95% CI: 1.07, 4.353] per
1-SD (1.56 (ig/dL) increase in erythrocyte Pb), but null associations with all B-cell NHL, diffuse large B-
cell lymphoma, CLL/SLL, other B-cell lymphoma, and MM.
10.4.7 Summary of Cancer Incidence and Mortality
The epidemiologic studies reviewed in the 2013 Pb ISA (U.S. EPA. 2013) reported inconsistent
findings across cancer endpoints. Among the studies that evaluated Pb exposure and overall cancer
incidence, there were no positive associations with various biological markers of Pb exposure. The
epidemiologic studies of overall cancer incidence were limited by the lack of biological measurements of
Pb and the lack of adjustment for potential confounders. The epidemiologic studies that examined the
associations between Pb concentrations and cancer mortality found inconsistent associations. Although
the studies were well-conducted longitudinal studies with control for a wide range of potential
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confounders, the studies were limited by the small number of cancer mortality cases, which reduces
statistical power to determine the presence of an association. The epidemiologic studies of Pb exposure
and lung cancer reported no evidence of an association. The studies available for review were conducted
in occupational cohorts and only included male study participants, which limits the generalizability of the
results. A few of the studies did not obtain Pb biomarker exposure levels or only used air sampling
measurements. Furthermore, these studies may be confounded by other workplace exposures and
covariates, such as smoking, that were not considered. There were a limited number of studies of brain
cancer and occupational Pb exposure. The associations between occupational Pb exposure and brain
cancer incidence and mortality varied depending on the tumor type or genetic variant. The implications of
the results from these studies were limited because they did not have individual-level biological Pb
measurements, relied on self-reported occupational exposure history, and did not control for potential
confounding by other workplace exposures. The epidemiologic studies reviewed relating to Pb exposure
and breast cancer suggested that women with breast cancer may have higher BLLs than those without
breast cancer. These studies were limited by their study designs, small sample sizes, and, with one study,
the method of Pb exposure measurement. There were also some inconsistent results among studies that
compared breast tissue Pb concentrations between breast tumor and control samples. The epidemiologic
literature reviewed for specific cancers and associations with Pb exposure reported varying associations
among occupational cohorts. Positive associations were observed between occupational exposure to Pb
and adenocarcinoma of the esophagus and stomach cancer, but there were inconsistent associations with
occupational Pb exposure and rectal cancer and occupational exposure to Pb in gasoline and stomach
cancer. These studies were limited to the study populations consisting of only men, no personal biological
or exposure measurements for Pb, and no control for potential confounding by other occupation
exposures.
While there were no recent PECOS-relevant epidemiologic studies of Pb exposure and overall
cancer incidence, lung cancer, and brain cancer, there were a limited number of recent epidemiologic
studies that examined the association between Pb concentrations and overall cancer mortality, breast
cancer mortality, and mortality from other cancers.
The recent PECOS-relevant epidemiologic studies reviewed were inconsistent across cancer
endpoints and support the conclusions from the 2013 Pb ISA (U.S. EPA. 2013). There were inconsistent
findings in large population-based studies examining the relationship between Pb exposure and overall
cancer mortality. While these recent epidemiologic studies were conducted in well-established cohorts,
the overall follow-up period was short (<11 years), there were a small number of cancer mortality cases
resulting in reduced precision across the studies, and there was a lack of control for some confounders
such as co-morbidities. Of note, the cohorts in the recent epidemiologic literature would generally be
expected to have had appreciable past exposures to Pb; however, the extent to which adult BLLs in these
cohorts reflect the higher exposure histories is unknown as to the extent to which these past Pb exposures
(magnitude, duration, frequency) may or may not elicit cancer incidence and/or mortality.
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Recent epidemiologic studies evaluating the associations between breast cancer and blood Pb
reported inconsistent findings, with an NHANES study finding increasing odds of breast cancer in higher
quartiles of blood Pb, while another study using three longitudinal cohorts in Italy, Sweden, and United
States did not find associations between breast cancer and blood Pb. The inconsistency in findings may be
related to difference in study design, biomarker of exposure, timing of exposure, range of Pb levels, and
difference in controlling for potential confounders (age at menarche, pregnancy history, oral contraceptive
use, female hormone use, and menopause status).
The recent epidemiologic literature for site-specific cancers and Pb exposure is limited, reporting
varied associations. The small body of evidence across various site-specific cancer endpoints limits the
ability to judge coherence and consistency across these studies, although the positive associations
reported demonstrate that Pb exposure could result in physiological responses that contribute to some
site-specific cancers. While these studies did control for a wide range of potential confounders, the studies
were limited by small number of cases, relatively short time between exposure and outcome, potential
differences in Pb exposure based on study location, and different biomarkers of exposure.
Overall, there were inconsistent findings in the limited number of epidemiologic studies assessing
associations between Pb exposure and cancer endpoints. While many of these studies utilized large
population-based cohorts, they were limited by the small number of cases, short follow-up time, range of
Pb levels, biomarkers of exposure, information of past Pb exposure, and lack of control of some potential
confounders.
10.5 Biological Plausibility
This section describes the biological pathways that potentially underlie cancer effects resulting
from exposure to Pb. Figure 10-1 graphically depicts these proposed pathways as a continuum of
pathophysiological responses—connected by arrows—that may ultimately lead to the apical cancer events
associated with exposures to Pb at concentrations observed in some epidemiologic studies (e.g., cancer
incidence and mortality). Most studies cited in this subsection are discussed in greater detail earlier in this
Appendix. Note that the structure of the biological plausibility sections and the role of biological
plausibility in contributing to the weight-of-evidence analysis used in the current Pb ISA are discussed in
Section 10.6.
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Altered binding
and transport
proteins, protein
confirmational
changes,
enhanced tissue
accumulation
r
Oxidative stress
Genotoxicity
Pb
and
and
Exposure
mitochondrial
| DNA damage and
dysfunction
repair
L
Epigenetic and
transcriptional
and translation
modifications
Note: The boxes above represent the effects for which there is experimental or epidemiologic evidence related to Pb exposure, and
the arrows indicate a proposed relationship between those effects. Solid arrows denote evidence of essentiality as provided, for
example, by an inhibitor of the pathway or a genetic knockout model used in an experimental study involving Pb exposure. Shading
around multiple boxes is used to denote a grouping of these effects. Arrows may connect individual boxes, groupings of boxes, and
individual boxes within groupings of boxes. Progression of effects is generally depicted from left to right and color-coded (gray,
exposure; green, initial effect; blue, intermediate effect; orange, effect at the population level or a key clinical effect). Here,
population level effects generally reflect results of epidemiologic studies. When there are gaps in the evidence, there are
complementary gaps in the figure and the accompanying text below. The structure of the biological plausibility sections and the role
of biological plausibility in contributing to the weight-of-evidence analysis used in the ISA are discussed in Section 10.6.
Figure 10-1 Potential biological pathways for cancer from exposure to Pb.
The development of cancer is a multistep process that involves the progressive accumulation of
mutations leading to upregulation of oncogenes and loss of function of tumor suppressor genes resulting
in uncontrolled cell growth and invasion of cancer cells within organ tissue. Pb is well-known to cause
cancers in animal models, however, the carcinogenic potential of Pb in humans is not well defined. As
discussed in the 2006 Pb AQCD, the ability of Pb to cause neoplastic transformation in human cells is
limited and is confounded by the fact that some studies utilize Pb chromate. Thus, observed effects may
be related to the effects of chromate as opposed to effects of Pb. Despite this, Pb possesses several
characteristics that were identified by the IARC that are common of human carcinogens (Smith et al..
2016). In addition, Pb is known to act on several pathways that could plausibly lead to cancer
development. The multifaceted pathway outlined in Figure 10-1 connects Pb exposure to cancer incidence
via Pb-protein binding, direct mutagenicity, genotoxicity, inflammation, oxidative stress, and epigenetic
changes. Together, the experimental evidence can provide plausibility for the carcinogenic potential
of Pb.
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The most direct pathway to Pb-induced carcinogenesis would involve mutagenesis in response to
Pb treatment that over time would result in cell transformation. As discussed in the 2006 Pb AQCD, there
is little evidence of the mutagenic potential of Pb (U.S. EPA, 2006). A recent study suggests that Pb can
directly interact with the DNA causing conformational change (Zhang et al., 2014). In this study Pb
caused increased markers of DNA damage although it is not clear if the binding of Pb was responsible for
the observed DNA damage. The potential for Pb to directly induce DNA mutations remains limited and,
as mentioned in the 2006 Pb AQCD, may only occur at very high concentrations.
The strongest data for potential carcinogenesis comes from experiments related to oxidative
stress-induced genotoxicity. The role of oxidative stress in the pathway of cancer is well documented
(Haves et al., 2020). Oxidative stress can result in the damage of proteins, lipids, and DNA. Pb exposure
is well known to cause oxidative stress in several organ systems. Oxidative stress is controlled by a
balance between the formation of reactive oxygen species (ROS) and the actions of antioxidant defenses.
As discussed in the 2013 Pb ISA, multiple in vitro experiments using diverse mammalian cell cultures
exposed to Pb compounds (Pb acetate, Pb chloride, Pb nitrate and divalent Pb ions) for different durations
result in increased production of ROS (U.S. EPA, 2013). This is supported by more recent studies that
consistently report increased ROS levels, decreased antioxidant defenses, and increased markers of
oxidative damage in Pb-exposed cells (see Section 10.3.2). The source of increased generation of ROS in
the context of cancer is not clear but could result as a byproduct of Pb-induced inflammation or Pb
displacement of biologically relevant ions in enzymes, especially those involved with metabolism and
energy production in the mitochondria.
Oxidative stress that damages DNA or impairs DNA repair can lead to mutation and subsequent
cellular transformation. As discussed in the 2013 ISA and in more recent studies, several markers of DNA
damage have been shown to be increased in Pb-exposed cells including 8-OH-deoxy guanine adducts (Liu
et al.. 20IS), alterations in comet DNA content, comet tail movement (Siddarth et al., 2018; El Makawy et
al., 2015; Shakoori and Ahmad, 2013), and DNA double strand breaks (as assessed by H2Ax foci) (Liu et
al., 2018; Shah et al., 2016; Pottier et al„ 2013) as well as diverse genotoxicity measures like micronuclei
formation (Martini et al., 2020; Alimba et al., 2016; Shah et al., 2016; El Makawy et al., 2015) and SCE
(Turkez et al., 2012). Similar increases in bone marrow micronuclei and increased comet tail movement
are seen in animal studies following Pb exposure (Olatunji-Ojo et al„ 2020; Okesola et al., 2019;
Nascimento and Martinez, 2016; El Makawy et al., 2015). In addition, the DNA repair rate has been
shown to be reduced in Pb treated cells (Martinez-Alfaro et al„ 2012). For example, the base excision
repair capacity of the DNA repair enzyme APE-1 is decreased by Pb treatment (Hernandez-Franco et al„
2018). Another study showed reduced DNA repair was associated with decreased glutathione suggesting
that oxidative stress might drive the reduction of DNA repair (Martinez-Alfaro et al., 2012). This data is
further bolstered by an experiment in humans exposed occupationally to Pb that show increased markers
of DNA damage and reduced DNA repair capacity (Jannuzzi and Alpertunga, 2016). In many
experimental cases, treatment with antioxidant compounds can protect against DNA damage (Okesola et
al., 2019; Siddarth et al„ 2018; El Makawy et al„ 2015) suggesting that oxidative stress is necessary for
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Pb-induced genotoxicity. This data supports a solid line in Figure 10-1 from oxidative stress to
genotoxicity.
Pb can also plausibly promote cancer development through induction of inflammation.
Inflammation is a hallmark of a pro-cancer environment. Induction of inflammation could be direct effect
by increased secretion of pro inflammatory markers. In addition, inflammation can result from cell
damage caused by oxidative stress. The 2013 Pb ISA and 2006 Pb AQCD discuss evidence that Pb
treatment can trigger the production of inflammatory mediators in vitro as well as in many organ systems
(U.S. EPA, 2013, 2006). More recent in vitro evidence supports these findings in the context of cancer
cell lines (Jiang et al., 2020; Lin et al., 2015). Many natural compounds that demonstrate anticancer
activity in vitro possess both anti-inflammatory and antioxidant capacity suggesting that inflammation
could be playing a role in the development of cancer.
Excessive DNA damage, as a result of inflammation and oxidative stress, can activate cell death
pathways. Cancer can arise when mutated cells suppress cell death pathways. Alternatively, cell death
often triggers compensatory expansion of surrounding cells. With chronic injury, a constant repair process
activation can trigger hyperplastic growth and degradation of extracellular matrix that can promote
cellular transformation and tumor invasiveness. While there is evidence that Pb treatment in vitro can lead
to cell death (see Section 10.3.5), there is no evidence to suggest that Pb can cause resistance to cell
death. However, there are some indications that Pb can stimulate cellular regrowth that over time could
potentially promote cellular transformation. Wang et al. (2013) showed that Pb treatment of CL3 cells
resulted in increased cell cycle progression. Another recent study showed that Pb treatment can lead to
increased MMP expression resulting in greater cell migration in a wound healing assay (Akin et al.,
2019). Together, there is strong evidence that Pb can cause cell death but the role of Pb in the
development of apoptosis resistance or uncontrolled cell growth remains speculative.
Over time, accumulation of mutations that promote tumor growth and blunt anti-tumor defenses
can lead to cell transformation and increased cancer incidence. In vitro assays can measure transformation
as an increase in morphologically distinct cells (i.e., a foci). As discussed in the 2013 ISA, data from
cellular transformation assays have shown that Pb acts as a promoter of cellular transformation in animal
cells in vitro. In support of this, a recent study showed that Pb pretreatment of Balb/c-3T3 cells prior to
transformation with n-methyl-n-nitrosoguanidine and 12-O-tetradecanoylphorbol-13-acetate resulted in
increased foci formation suggesting that Pb can help to promote transformation (Hernandez-Franco et al.,
2018).
Changes in regulation of gene expression through epigenetic mechanisms represent another
plausible pathway by which Pb can promote tumor formation. The 2013 ISA provided limited evidence
from human studies that tibia Pb levels could be inversely related to global methylation markers (U.S.
EPA, 2013). A new study of infant blood spots showed a general decrease in methylation at 33 CpG sites
with increasing BLLs (Laurino et al., 2020). Interestingly, pathway enrichment analysis suggested that
differentially methylated sites corresponded to cell morphogenesis and cell adhesion. This suggests that
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changes in epigenetics regulation could play a role in changes in cell adhesion which could be important
in the context of tumor invasiveness and metastasis. Increased methylation was also seen in the promoter
regions of several DNA repair genes following Pb exposure which correlated with decreased repair
protein levels (Liu et al.. 2018). Alterations of methyltransferases levels following Pb exposure in vitro
has also been reported and correlate with increased expression of an oncogene (Ghosh et al.. 2018).
Insight into the mechanism of epigenetic regulation by Pb was provided by Rabbani-Chadcgani et al.
(2011) who showed that Pb nitrate bound to rat liver chromatin. When analyzed separately, Pb bound
histones with higher affinity than to DNA (Rabbani-Chadcgani et al.. 2011). The affinity of Pb nitrate was
greater than Ni nitrate in these studies. Though the biological effects of histone binding were not
investigated, it is possible that binding of Pb to histone chromatin or histones could result in epigenetics
changes through alterations in accessibility of DNA or histones to modifying enzymes. Overall, there is
evidence that Pb can affect epigenetic markers of genes that could affect cancer development.
Pb has been shown to replace biologically relevant ions within cellular proteins which can cause
confirmational changes that can impair target protein function. Thus, direct binding of Pb to cellular
proteins could form another plausible pathway to promote tumor formation. For example, Pb can compete
with Zn in Zn finger domains which are present in several transcription factors (Ghering et al.. 2005;
Huang et al.. 2004; Hanas et al.. 1999). Pb-induced conformation changes in cellular proteins could have
widespread effects on cellular functions and could theoretically promote cellular transformation. The
potential of Pb to directly bind and alter cellular protein function represents another pathway by which Pb
exposure could result in cell transformation and tumorigenesis.
Together, mechanistic toxicological data provides several possible pathways through which Pb
exposure can result in the tumorigenesis that is seen in animal studies and that is reported in some
epidemiologic studies. The evidence is strongest for a pathway that involves Pb-induced inflammation
and oxidative stress which causes subsequent DNA damage that, in conjunction with suppression of
proper DNA repair mechanisms, can lead to mutations that could result in neoplastic transformation.
There is also increasing evidence for the plausibility of epigenetic changes caused by Pb to promote
tumorigenesis. Given the widespread impacts of Pb on cellular proteins there are other plausible pathways
for tumor formation including direct mutagenesis and chronic tissue damage with subsequent cell cycle
disruption, although the evidence for these pathways is more limited.
10.6 Summary and Causality Determination
The 2013 Pb ISA concluded that there was a "likely to be a causal relationship" between Pb
exposure and cancer (U.S. EPA. 2013). This causality determination was made on the basis that the
toxicological literature provides consistent evidence of the carcinogenic potential of Pb and possible
contributing modes of action, including genotoxic, mutagenic, and epigenetic effects. The toxicological
literature provided strong evidence for cancer following long-term exposure (i.e., 18 months or 2 years) to
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high concentrations of Pb (>2,6000 ppm) in drinking water. The consistent evidence indicating Pb-
induced carcinogenicity in animal models was substantiated by findings from multiple high-quality
toxicological studies in animal and in vitro models from different laboratories. Carcinogenicity in animal
toxicology studies with relevant routes of Pb exposure has been reported in the kidneys, testes, brain,
adrenals, prostate, pituitary, and mammary gland, albeit at high doses of Pb. Epidemiologic studies of
cancer incidence and mortality reported inconsistent results; one strong epidemiologic study demonstrated
an association between blood Pb and increased cancer mortality (Schober et al.. 2006). but the other
studies reported weak or no associations (Khalil et al.. 2009; Weisskopf et al.. 2009; Menke et al.. 2006).
Although there are no recent PECOS-relevant animal toxicological studies evaluating the
relationship between Pb exposure and cancer endpoints, the animal studies available in previous reviews
continue to provide strong support for the carcinogenic potential of high Pb exposures (chronic
10,000 ppm Pb acetate diet or 2,600 ppm drinking water Pb acetate) (Tokar et al.. 2010; Waalkes et al..
1995; Kasprzak et al.. 1985; Koller et al.. 1985; Azar et al.. 1973; Van Esch and Kroes. 1969). Recent in
vitro studies report Pb activation of pathways that are relevant and frequently reported to be involved in
cancer development and/or progression, particularly pathways mediated by oxidative stress, genotoxicity,
and inflammation. Other mechanistic pathways that may be involved in Pb-induced carcinogenesis
include cell cycle regulatory genes, epigenetics, apoptosis, and necrosis with predictive regenerative
proliferation. Additionally, new areas of research involving MMPs and metallothionines have emerged
and provide evidence of other potential mechanistic pathways through which Pb exposure could
contribute to cancer. This recent evidence has added to our understanding of how Pb exposures may
activate the mechanistic pathways that can result in cancer.
Recent epidemiologic studies that examined the associations between Pb exposure and overall
cancer mortality reported inconsistent results, similar to the epidemiologic studies evaluated in the 2013
Pb ISA (U.S. EPA. 2013). The recent studies of overall cancer mortality used exposure data from
population-based national surveys linked to mortality records. While there were positive associations
between blood Pb and overall cancer mortality in large population survey studies in the United States and
Korea (Bvun et al.. 2020; Duan et al.. 2020). there were null associations in another NHANES study (van
Bemmel et al.. 2011). These epidemiologic studies were conducted in large, well-established population-
based cohorts, but there are still limitations. These include short overall follow-up periods (<11 years), a
small number of cancer mortality cases resulting in reduced precision across the studies, and a lack of
control of some confounders such as co-morbidities. There were a limited number of recent
epidemiologic studies evaluating the associations between Pb exposure and site-specific cancers. The
studies reviewed reported inconsistent findings. While several of the studies were well-conducted in large
cohorts, there remain uncertainties in the biomarkers of exposure (blood versus erythrocytes), timing of
exposure, years of follow-up, range of Pb levels, exposure circumstances (magnitude, duration, timing,
and frequency) and differences in controlling for potential confounders (co-morbidities, BMI, age at
menarche, pregnancy history, oral contraceptive use, female hormone use, and menopause status).
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In summary, the collective body of evidence is sufficient to conclude that there is likely to be
a causal relationship between Pb exposure and cancer. The key evidence for this causal determination
is in Table 10-1. There continues to be strong evidence from in vivo toxicological studies and from
studies of mechanistic pathways indicating the carcinogenic potential of Pb exposure, including
inflammation; oxidative stress; and genotoxic, mutagenic, and epigenetic effects. Recent mechanistic
research further identifies biologically plausible molecular pathways through which Pb could contribute
to the initiation and/or progression of cancer, and these pathways are consistent with the IARC 10 key
characteristics of carcinogenic mechanistic pathways (Smith et al.. 2016). Several of these pathways are
consistent with the reported mechanistic pathways associated with Pb carcinogenicity reported in the
2013 Pb ISA. Recent epidemiologic studies provide inconsistent evidence of associations between Pb
exposure and cancer incidence and/or mortality, for either overall or site-specific cancer. More
specifically, the small body of epidemiologic evidence across various site-specific cancer endpoints limits
the ability to judge coherence and consistency across these studies, although the positive associations
observed in a small number of studies at relevant BLLs demonstrate that Pb exposure could result in
physiological responses that contribute to urothelial carcinoma, gastrointestinal cancer, non-Hodgkin's
lymphoma, and multiple myeloma. Despite uncertainty due to inconsistent findings across epidemiologic
studies, animal toxicology studies and in vitro mechanistic studies provide strong evidence for the
carcinogenic potential of Pb exposures.
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Table 10-1 Summary of evidence for a likely to be causal relationship between Pb exposure and cancer
RatiDet5m°naCtionaality KeV Evidence' Key References' Pb Biomarker Levels Associated with
Consistent evidence from Consistent findings across multiple Azar et al. (1973)
multiple animal studies with
chronic Pb exposure
toxicology studies using 18-mo or2-yr
cancer bioassays in rats wherein
rodents are fed chow or received
drinking water enriched with Pb
acetate and show tumor development.
Kasprzak et al. (1985)
Kolleret al. (1985)
Van Esch and Kroes (1969)
See Section 10.3.2
Chronic 10,000 ppm Pb acetate diet or
2,600 ppm drinking water Pb acetate,
no blood Pb measurement available.
Gestational and lactational Pb
exposure induced carcinogenicity in
adult offspring.
Waalkes et al. (1995)
Tokaret al. (2010)
See Section 10.3.2
500, 750, and 1,000 ppm Pb in drinking
water, no blood Pb measurement
available.
Most evidence clearly
supports biological
plausibility
Consistent toxicological evidence for
mutagenicity, carcinogenicity, and
genotoxicity of Pb reported by multiple
laboratories in humans, animals and in
vitro models using multiple assays
(micronuclei, SCE, comet).
See subsections in Section 10.3
Toxicology evidence of DNA and cellular
damage:
Tapisso et al. (2009)
Alqhazal et al. (2008)
Gastaldo et al. (2007)
Xu et al. (2008)
Nava-Hernandez et al. (2009)
Yediou et al. (2010)
Xu et al. (2006)
Kermani et al. (2008)
Epidemiology evidence of DNA and cellular
damage:
Wiwanitkit et al. (2008)
Duvdu et al. (2005)
Khan etal. (2010)
Olewihska et al. (2010)
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Rationale for Causality
Determination3
Key Evidence"
Key Referencesb
Pb Biomarker Levels Associated with
Effects0
Some evidence for epigenetic
changes. Bone Pb levels were
inversely associated with LINE-1
methylation in a study of adult men
Shaik and Jamil (2009)
Wright etal. (2010)
Patel (2013)
Study showed inverse association
between maternal postpartum bone Pb
levels and Alu and LINE-1 methylation Pilsner etal. (2009)
in cord blood.
Occupational battery workers had
ALAD hypermethylation compared with
controls; cell culture study of high dose Li et al. (2011)
Pb exposure caused ALAD
hypermethylation.
Toxicological evidence of
clastogenic (SCE,
micronucleus formation,
chromosomal aberrations),
mutagenic, and genotoxic
effects with Pb chromate
Wise etal. (2010)
Grlickova-Duzevik et al. (2006)
Saverv et al. (2007)
Camvre et al. (2007)
Stackpole et al. (2007)
Li Chen et al. (2009)
Wise et al. (2009)
Wise etal. (2011)
Some toxicological studies employ Pb
chromate when investigating the
clastogenic, mutagenic, and genotoxic
effects of Pb. The effect of the
chromate ion in contributing to these
effects cannot be ruled out.
Holmes et al. (2006a)
Wise et al. (2006a)
Holmes et al. (2006b)
Wise et al. (2006b)
Yipptai r?nrm
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Rationale for Causality
Determination3
Key Evidence"
Key Referencesb
Pb Biomarker Levels Associated with
Effects0
Epidemiologic evidence is
limited and inconsistent
Epidemiologic studies of overall cancer
mortality have inconsistent findings.
These are high-quality, longitudinal
studies and control for potential
confounders, such as age, smoking,
and education. The follow-up period
was short (<11 yr). There is uncertainty
related to exposure patterns resulting
in likely higher past Pb exposure.
There was the lack of control of
potential important confounders such
as co-morbidities.
Overall Cancer Mortality: See Section 10.4.2
In the mortality studies, the majority of
the study participants' BLLs were <10 |jg/
dL (NHANES medians ranged from 1.49
to 7.5 |jg/dL and KNHANES geometric
mean was 2.26 |jg/dL).
Epidemiologic studies of specific
cancer sites were limited. Many of the
epidemiologic studies examining
specific cancer sites were case-control
studies and not all included potentially
important confounders, such as
smoking and co-morbidities. There is
uncertainty related to exposure
patterns resulting in likely higher past
Pb exposure and impact of difference
biomarkers (blood vs. stored
erythrocytes).
Specific Cancer:
Breast Cancer: See Section 10.4.5
Other Cancer: See Section 10.4.6
In studies of breast cancer, the majority
of the study participants' BLLs were
<10 ug/dL (medians ranged from 1.15-
8.78 pg/dL).
In studies of other cancer, the majority of
the study participants' BLLs were
<10 pg/dL (medians ranged from 3.05-
9.191 pg/dL and means ranged from
2.56-2.81 pg/dL).
aBased on aspects considered in judgments of causality and weight of evidence in causal framework in Table I and Table II of the Preamble to the ISAs (U.S. EPA. 2015).
bDescribes the key evidence and references, supporting or contradicting, contributing most heavily to causality determination and, where applicable, to uncertainties or
inconsistencies. References to earlier sections indicate where the full body of evidence is described.
°Describes the Pb biomarker levels at which the evidence is substantiated.
ALAD = 6-aminolevulinic acid dehydratase; BLL = blood lead level; KNHANES = Korea National Health and Nutrition Examination Survey; LINE = long interspersed nuclear
elements; mo = month; NHANES = National Health and Nutrition Examination Survey; Pb = lead; SCE = sister chromatid exchange; yr = year.
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10.7 Evidence Inventories - Data Tables to Summarize Study Details
Table 10-2 Epidemiologic studies of exposure to Pb and cancer effects
Reference and
Study Design Study PoPulatlon
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Overall Cancer Mortality
Menke et al.
{2006)t
U.S.
NHANES III (1988-
1994), mortality
follow-up in 2001
(12 yr follow-up)
Cohort
NHANES III
n = 13,946, >20 yr
Blood
Blood was measured by
GFAAS with Zeeman
correction
Mean: 2.58 |jg/dL
Blood Pb Tertiles:
T1: <1.93 |jg/dL
T2: 1.94-3.62 pg/dL
T3: >3.63 pg/dL
Age of Measurement
Mean 44.4 yr
Overall cancer mortality
Cause of death was
determined by the underlying
cause of death listed on
death certificates. ICD-9
codes (codes 140 to 239)
were used for deaths
between 1988 and 1998 and
ICD-10 codes (C00-C97 and
D00-D048) were used for
deaths during 1999 and
2000.
Age at Outcome: Age at
death
Cox proportional
hazard regression
analysis adjusted
age, race/ethnicity,
sex, urban residence,
cigarette smoking,
alcohol consumption,
education, physical
activity, household
income, menopausal
status, BMI, CRP,
total cholesterol,
diabetes mellitus,
hypertension, GFR
category
HR:
T1
T2
T3
Reference
0.72 (0.46, 1.12)
1.10 (0.82, 1.47)
Schober et al.
(2006)1-
U.S.
NHANES III (1988-
1994), mortality
follow-up in 2006
-8.55 yr of follow-
up
NHANES III
n = 9,686, >40 yr of
age
Blood
Blood was measured by
GFAAS with Zeeman
correction
Age of Measurement:
>40 yr
Blood Pb Tertiles:
Overall cancer mortality
Deaths due to malignant
neoplasm (ICD-10 codes
C00-C97)
Age at Outcome: Age at
death
Cox proportional
hazard regression
analysis adjusted for
sex, age,
race/ethnicity,
smoking, education
level
Relative Risk (RR):
T1: Reference
T2: 1.44 (1.12, 1.86)
T3: 1.69 (1.14, 2.52)
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Cohort
T1 < 5 (median 2.6 |jg/dL)
T2 5-9 (median 6.3 |jg/dL)
T3 > 10 (median
11.8 |jg/dL)
van Bemmel et al.
(2011)
U.S.
NHANES III (1984-
1994), mortality
follow-up in 2007
(-7.8 yr of follow-up
for low blood Pb
and -7.5 yr of
follow-up for high
blood Pb)
Cohort
NHANES
n: 3,349 (BLL <5 pg/dL
n: 2,532; BLL >5 ug/dL
n: 817)
NHANES III (1984-
1994) general
population restricted to
the participants who
were successfully
genotyped, excluding
those under the age of
40; those with no
baseline blood Pb
measurements;
missing data on ALAD
genotype, education,
and date of study entry
Blood
Blood was measured by
GFAAS
Age at Measurement:
40+
Median for BLL <5 pg/dL:
2.6 pg/dL
Median for BLL >5 pg/dL:
7.5 pg/dL
Max: 52.9 pg/dL
Overall cancer mortality
Mortality from malignant
neoplasm (ICD-10 codes
C00-C97)
Age at Outcome:
Age at death was defined as
the time to event
Cox proportional
hazard regression
models were adjusted
for age, education,
sex, smoking status,
race/ethnicity, ALAD
genotype
HR: 1.08 (0.98, 1.19) for BLL
>5 pg/dL, compared to
<5 pg/dL
HR for ALADGG: 1.08 (0.99,
1.19) for BLL >5 pg/dL,
compared to <5 pg/dL
Duan et al. (2020)
U.S.
1999-2014,
mortality follow-up
in 2015 (-7.1 yr of
follow-up)
Cohort
NHANES
n: 26,056
NHANES participants
aged 20 yr or older,
not pregnant, or
missing covariate data
Blood
Blood was measured by
multielement atomic
absorption spectrometer
with a Zeeman background
correction (NHANES 1999-
2002) or ICP-MS (after
2003)
Age at Measurement:
average age: 45.9 yr
Medianb: 1.49 pg/dL
75thb: 2.31 pg/dL
Overall cancer mortality
Death certificates were used
to determine the source and
cause of death, specifically
cancer-specific mortality
(codes C00-C97)
Age at Outcome:
Age at death
Poisson regression
models estimated the
RR and adjusted for
sex, age, age
squared, and ethnicity
(Model 1); plus
education, PIR,
cotinine category,
BMI, and physical
activity (Model 2);
plus hypertension and
diabetes (Model 3)
RR per one unit increase in
blood Pb
Model
1:
1
65
(1
38,
1
97)
Model
2:
1
47
(1
22,
1
77)
Model
3:
1
47
(1
22,
1
78)
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Bvun et al. (2020)
Korea
2007-2015,
mortality follow-up
in 2018 (between 3
and 11 yr of follow-
up)
Cohort
KNHANES
n: 7,308
Individuals with a BLL
less than 10 |jg/dL,
who were aged 30 yr
and over at the
baseline examination,
and who were not
diagnosed with cancer
or ischemic heart
disease
Blood
Blood was measured by
GFAAS with Zeeman
background correction
Age at Measurement:
30+ yr
Geometric mean: 2.26
(±1.52) |jg/dL
Blood Pb tertiles:
T1: <1.91 |jg/dL
T2: 1.91-2.71 pg/dL
T3: >2.71 pg/dL
Overall cancer mortality
Deaths identified from all
non-accidental causes (the
International Classification of
Disease tenth revision: ICD-
10, A00-R99) and cancer
(ICD-10, C00-97).
Age at Outcome:
Age at death
Cox proportional
hazard models: Initial
models (Model 1)
were adjusted only for
age and sex.
Subsequent models
(Model 2) were
additionally adjusted
for household
income, education,
occupation, smoking
status, drinking
frequency, BMI, and
physical activity. Final
models (Model 3)
were further adjusted
for intake of high-
lead-containing food
intake (grains,
vegetables, and
seafood).
HR
Model 1:
Reference
3.19 (1.47, 6.91)
2.41 (1.17, 4.96)
Model 2:
Reference
3.46 (1.65, 7.26)
2.26 (1.09, 4.69)
Model 3:
Reference
3.42 (1.65, 7.08)
2.27 (1.09, 4.70)
Breast Cancer
Gaudetetal. (2019)
United States,
Sweden
Italy,
CPS-II n: 21,956;
EPIC-ltaly n: 32,578;
NSHDS n: 40,256
U.S. CPS-II: 1998-
2001; EPIC-ltaly:
1993-1998;
NSHDS: 1990-
2006
Cohort
Blood (erythrocytes)
Blood was measured by
ICP-MS
Age at measurement:
Median age (range): CPS-
II: 68 (47-85); EPIC-ltaly:
52 (35-70); NSHDS: 50
(30-61)
Median0: CPS-II:
2.53 pg/dL; EPIC-ltaly:
8.78 pg/dL; NSHDS:
3.897 pg/dL
Breast cancer
CPS-II: Cancer incident to
blood draw diagnosed
through June 30, 2011 were
self-reported on follow-up
questionnaires and
subsequently verified by
obtaining medical records or
through linkage with state
registries when complete
medical records could not be
obtained. Deaths were
obtained through linkage of
the cohort with the National
Death Index.
Logistic regression
models estimated the
relative risk (RR);
adjusted for race,
blood draw date and
age for CPS-II; and
age, year of blood
collection,
menopausal status
and Italian study
center for EPIC-ltaly
and NSHDS
CPS-II:
RR per each unit increase in
blood Pb (continuous): 1.00
(0.99, 1.00)
Quintile RR:
Q1
Reference
Q2
00
o
1.49)
Q3
1.07 (0.79,
1.45)
Q4
0.94 (0.69,
1.28)
Q5
0.94 (0.69,
1.28)
EPIC-ltaly:
10-30
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Breast cancer cases
included 816 cases
from CPS-II, 294 from
EPIC-ltaly and 325
from NSHDS. Each
case was paired with
one control. Eligible
controls were selected
among those who
were alive and cancer-
free at the time of the
case's diagnosis and
matched on race
(CPS-II), birthdate
(within 6 mo in CPS-II
and within 2.5 yr in
EPIC-ltaly and
NSHDS), menopausal
status (NSHDS, EPIC-
ltaly), study center
(EPIC-ltaly) and blood
draw date (within 6 mo
in CPS-II and within
the same year in
EPIC-ltaly and
NSHDS).
75thc: CPS-II: 3.442 pg/dL;
EPIC-ltaly: 11.21 pg/dL;
NSHDS: 5.288 pg/dL
Blood Pb Quintiles0:
CPS-II:
Q1
0-1.68 pg/dL
Q2
1.69-2.28 pg/dL
Q3
2.29-2.88 pg/dL
Q4
2.89-3.76 pg/dL
Q5
3.77-14.84 pg/dL
EPIC-ltaly:
Q1
2.40-6.35 pg/dL
CM
a
6.36-7.99 pg/dL
Q3
8.00-9.99 pg/dL
Q4
10.00-12.50 pg/dL
Q5
12.51-39.18 pg/dL
NSHDS:
Q1
0.80-2.64 pg/dL
Q2
2.65-3.57 pg/dL
Q3
3.58-4.54 pg/dL
Q4
4.55-5.53 pg/dL
Q5
5.54-22.37 pg/dL
EPIC-ltaly: Newly identified
cancer cases were identified
through automated linkages
to cancer and mortality
registries, municipal
population offices and
hospital discharge systems.
In Naples, follow-up
information was collected
through periodic personal
contact.
NSHDS: Newly identified
cancer cases were identified
through linkage with the
Swedish Cancer Registry and
the local Northern Sweden
Cancer Registry.
Age at Outcome: Age at
diagnosis
RR per each unit increase in
blood Pb (continuous): 1.00
(0.99, 1.00)
Quintile RR:
Q1
Q2
Q3
Q4
Q5
Reference
0.94 (0.57, 1.56)
0.96 (0.57, 1.61)
0.74 (0.43, 1.25)
0.77 (0.45, 1.33)
NSDHS:
RR per each unit increase in
blood Pb: 1.00 (0.99, 1.01)
Quintile RR:
Q1
Q2
Q3
Q4
Q5
Reference
1.09 (0.68, 1.76)
0.99 (0.61, 1.60)
0.65 (0.39, 1.08)
1.06 (0.66, 1.71)
10-31
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Wei and Zhu (2020) NHANES
n: 9,260
U.S.
2003-2012 Female participants
20 yr of age or older
Cross-sectional
Blood
Blood was measured by
ICP-MS
Age at measurement:
20+ yr
Geometric
mean:1.09 |jg/dL
Median: 1.15 |jg/dL
Max: 25 |jg/dL
Blood Pb Quartiles:
Q1
Q2
Q3
Q4: >1.8 |jg/dL
<0.8 |jg/dL
0.8 to <1.2 |jg/dL
1.2 to <1.8 |jg/dL
Breast cancer
Self-reported cancer
diagnosis was obtained from
the medical conditions
questionnaires. Participants
were being asked a question
"Have you ever been told by
a doctor or other health
professional that you had
cancer or a malignancy of
any kind?'. Participants who
answered "yes" were
subsequently asked "What
kind of cancer was it? Only
women who reported "no
cancer" diagnosis or a
"breast cancer" diagnosis
were included in our study
population. The study
population was categorized
into with breast cancer and
without breast cancer in the
analytical models.
Age at Outcome:
age at diagnosis
Logistic regression
models were adjusted
for age, race/ethnicity,
poverty status,
education, BMI,
physical activity, age
at menarche,
pregnancy history,
oral contraceptive
use, female hormone
use, cigarette
smoking, and alcohol
consumption
OR
Q1
Q2
Q3
Q4
Reference
2.52 (1.35, 4.73)
2.01 (1.05,
2.63 (1.36,
3.84)
5.09)
10-32
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Other Cancers
Chung et al. (2017)
Taichung
Taiwan
June 2011-August
2014
Case-control
n: 209 patients with
UC and 417 control
patients
UC patients aged 26-
96 yr, whose
diagnoses were
evaluated by a
pathologist. Matched
control participants
with cases according
to gender and age
(±3 yr) from patients
undergoing adult
health examinations.
Blood
Blood was measured by
ICP-MS
Age at Measurement:
Mean age for cases:
67.18 ± 10.79; mean age
for controls: 66.20 ± 10.06
Mean for cases: 2.81 |jg/dL
Mean for controls:
2.56 |jg/dL
Blood Pb Quartiles:
Q1: <1.76 |jg/dL
Q2: 1.76-2.31 pg/dL
Q3: 2.31-2.99 pg/dL
Q4: >2.99 pg/dL
Other cancers: Urothelial
carcinoma
Patients with UC comprised
outpatients or inpatients
among men and women
aged 30-90 yr old; UC cases
were limited to patients with
urinary tract urothelial
carcinoma, whose diagnoses
were evaluated by a
pathologist.
Age at Outcome:
Mean age for cases:
67.18 ± 10.79; mean age for
controls: 66.20 ± 10.06
Logistic regression
models were adjusted
for age, gender,
smoking
OR
Q1
Q2
Q3
Q4
Reference
0.68 (0.40, 1.15)
1.05 (0.64, 1.70)
1.66 (1.05, 2.61)
OR for smokers:
T1
T2
T3
Reference
1.71 (0.63, 4.60)
1.76 (0.69, 4.46)
OR for non-smokers:
T1
T2
T3
Reference
0.72 (0.43, 1.22)
1.40 (0.91, 2.39)
Blood Pb Tertiles for
Smoking Status:
T1: <1.98 pg/dL
T2: 1.98-2.73 pg/dL
T3: >2.73 pg/dL
10-33
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Lin etal. (2018)
Chaoshan,
China
June-December
2014
Case-control
n: 180 cases and 120
controls
Participants recruited
were native inhabitants
living in the Chaoshan
area (including the
cities of Shantou,
Chaozhou, and
Jieyang, and other
neighboring areas).
Cases and controls
had no distinction
between geographic or
cultural groups since
they were the native
aborigines in
Chaoshan.
Blood
Blood was measured by
GFAAS
Age at Measurement:
Cases mean age: 59.065;
Controls mean age: 47.09
Median0 for cases:
6.003 |jg/dL
Median0 for controls:
5.384 |jg/dL
75th° for Cases:
9.086 |jg/dL
75thc for Controls:
7.627 |jg/dL
Blood Pb Quartiles:
Q1: <25th percentile
Q2: 25th-50th percentile
Q3: 50th-75th percentile
Q4: >75th percentile
Other cancers:
Gastrointestinal cancers
All cases were newly
diagnosed and previously
untreated. Clinical
characteristics, including
basic medical data, were
obtained from medical
records. Controls (n = 112)
were recruited and found no
disease in the subsequent B-
ultrasound, imaging
examination, and
hematological examination.
Age at Outcome:
Cases mean age: 59.065;
Controls mean age: 47.09
Logistic regression
models were adjusted
for gender, age,
tobacco smoking, and
alcohol drinking
OR
Q1
Q2
Q3
Q4
Reference
0.683 (0.328, 1.423)
0.865 (0.410, 1.822)
2.32 (1.01, 4.94)
OR for Clinical Stages:
I: Reference
II: 2.099 (0.451, 9.759)
III: 1.458 (0.419, 5.074)
IV: 0.613 (0.210, 1.789)
OR for T Classification:
T1+T2: Reference
T3+T4: 4.752 (1.299, 17.389)
OR for N Classification:
NO: Reference
N1+N2+N3: 3.000 (0.822,
10.945)
OR for M Classification:
M0: Reference
M1: 4.546 (0.757, 27.317)
10-34
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Kelly etal. (2013)
Italy and Sweden
Italy: 1993-1998;
Sweden: 1990-
2006
Case-control
E n vi roGe n o M a rke rs
Study
n: Italy: n = 47,749;
Sweden: n = 95,000
The
E n vi roGe n o M a rke rs
study is based on
participants from two
existing prospective
cohort studies: EPIC-
Italy and the NSHDS.
EPIC-ltaly: 47,749
volunteers aged 35-
70 yr were enrolled in
five participating
centers across Italy.
The NSHDS includes
participants from the
Vasterbotten. A total of
95,000 healthy
individuals aged 40-60
were invited for
inclusion in the project
between 1990 and
2006.
Blood (erythrocytes)
Blood was measured by
ICP-MS
Age at Measurement:
Mean age for cases:
53.08 yr
Mean age for controls:
53.09 yr
Median0: 9.191 |jg/dL in
Italy
Median0: 4.499 |jg/dL in
Sweden
Erythrocyte Pb Quartiles0
for B-cell NHL:
Q1
Q2
Q3
Q4
1.5423-3.9286 pg/dL
3.9504-5.8763 pg/dL
5.8832-8.7218 pg/dL
8.7531-40.0843 pg/dL
Erythrocyte Pb Quartiles0
for B-cell NHL for Males:
Q1
Q2
Q3
Q4
Other cancers: B-cell non-
Hodgkin's lymphoma and
multiple myeloma
Lymphoma cases that
occurred within the two
cohorts between 2 and 16 yr
of follow up were identified.
Lymphoma cases were
classified into subtypes
according to the SEER
ICD-0-3 morphology codes.
All eligible B-cell NHL cases,
including multiple myeloma
were included.
Age at Outcome:
mean age for cases: 53.08 yr
mean age for controls:
53.09 yr
Conditional logistic
regression models
were adjusted for sex,
age, center, batch
and sample date
1.5423-4.4989 pg/dL
4.5444-6.1498 pg/dL
6.1904-10.0201 pg/dL
10.0528-
37.8943 pg/dL
Erythrocyte Pb Quartiles0
for B-cell NHL for Females:
Q1: 1.7019-3.6079 pg/dL
Q2: 3.6719-5.4739 pg/dL
OR:
B-cell NHL:
Total study population
Q1: Reference
Q2: 0.93 (0.51, 1.67)
Q3: 0.91 (0.47, 1.79)
Q4: 0.93 (0.43, 2.02)
p for trend: 0.849
Males:
Q1: Reference
Q2: 0.57 (0.23, 1.37)
Q3: 0.83 (0.35, 1.99)
Q4: 0.74 (0.27, 2.04)
p for trend: 0.742
Females:
Q1
Q2
Q3
Q4
Reference
0.62 (0.23, 1.65)
0.54 (0.20, 1.46)
0.42 (0.12, 1.47)
p for trend: 0.17
MM:
Total study population:
Q1: Reference
Q2: 1.30 (0.44, 3.86)
Q3: 1.17 (0.38, 3.59)
Q4: 1.63 (0.45, 5.94)
p for trend: 0.533
Males:
Insufficient data for models
Females:
Q1: Reference
10-35
-------
Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
Q3: 5.5401-7.7823 pg/dL
Q2:
0.71 (0.20,
2.57)
Q4: 7.8313-40.0843 pg/dL
Q3:
0.71 (0.19,
2.61)
Q4:
0.74 (0.14,
3.83)
Erythrocyte Pb Quartiles0
for MM:
Q1
Q2
Q3
Q4
1.1199-3.5133 [jg/dL
3.5184-5.1973 pg/dL
5.2459-7.9079 [jg/dL
8.1448-67.2484 pg/dL
Erythrocyte Pb Quartiles0
for MM for Males:
Q1
Q2
Q3
Q4
1.9898-3.6049 [jg/dL
3.8613-5.2578 pg/dL
5.2808-9.3128 pg/dL
9.7683-67.2482 pg/dL
Erythrocyte Pb Quartiles0
for MM for Females:
Q1
Q2
Q3
Q4
1.1199-3.0604 pg/dL
3.2928-4.8623 pg/dL
4.9859-7.5424 pg/dL
7.6344-22.0943 pg/dL
p for trend: 0.692
OR by NHL subtype
associated with a one unit
increase in log transformed
exposure levelsd:
MM:
Total study population: 1.04
(0.57, 1.90)
Males: 0.83 (0.35, 1.96)
Females: 1.28 (0.53, 3.08)
DLBCL:
Total study population: 0.60
(0.26, 1.40)
Males: 0.97 (0.35, 2.64)
Females: 0.29 (0.07, 1.18)
B-cell CLL:
Total study population:
0.71 (0.32, 1.57)
Males: 0.63 (0.23, 1.74)
Females: 0.79 (0.17, 3.60)
FL:
Total study population:
1.17 (0.52, 2.63)
Males: 0.80 (0.25, 2.55)
Females: 1.91 (0.54, 6.78)
10-36
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
OR for BLL >10 pg/dL:
B-cell NHL:
Total study population: 1.10
(0.60, 2.02)
Males: 0.93 (0.44, 1.98)
Females: 1.50 (0.53, 4.21)
MM:
Total study population: 1.29
(0.48, 3.45)
Males: 0.80 (0.21, 2.98)
Females: 2.50 (0.49, 12.89)
Deubler et al.
(2020)
U.S.
1992-1993 (1998-
2001)
Case-control
Cancer Prevention
Study-ll Nutrition
Cohort (CPS-II NC)
n: 375 B-cell NHL or
MM cases (95 DBLCL,
90 CLL/SLL, 62 FL, 76
MM and 52 other
B-cell lymphoma) and
750 matched controls
The CPS-II NC was
initiated in 1992 to
1993 and enrolled
184,185 men and
women aged 40 to 90
(median = 62.0) yr
residing in 21 states.
Participants self-
reported exposure
information and cancer
diagnoses by
completing an initial
baseline questionnaire
in 1992 to 1993 and
biennial follow-up
Blood (erythrocytes)
Blood was measured by
ICP-MS
Age at Measurement:
Average age of cases at
the time of blood draw:
69.8 yr
Average age of controls at
time of blood draw: 69.9 yr
Median0: 3.05 pg/dL
Maxc: 13.88 pg/dL
Erythrocyte Pb Quartiles0:
Entire cohort:
Q1: Oto <2.1008 pg/dL
Q2: 2.1008 to
<3.0268 pg/dL
Q3: 3.0268 to <4.094 pg/dL
Q4: >4.094 pg/dL
Other cancers: B-cell NHL
and multiple myeloma
Self-reported cancer
diagnoses were verified
through medical records or
state cancer registry linkage.
Verified incident B-cell NHL
(B-NHL) and MM were
identified from CPS-II NC
participants who were
cancer-free at time of blood
collection (1998 and 2001).
B-NHL cases were further
categorized into the following
subtypes using the 2008
WHO classification scheme:
CLL/SLL, DLBCL, FL, MM,
and other B-cell lymphoma.
Age at Outcome:
Average age at diagnosis:
75 yr
Conditional logistic
regression models
estimated relative
risks (RR) adjusted
for smoking status
(current, former,
never), average
alcohol consumption
(none, <1,1, >2,
missing drinks per
day) and multivitamin
use in the week prior
to blood draw (yes,
no, missing), based
on a 10% change in
the parameter
estimates criterion
RR:
Overall lymphoid malignancy
Entire cohort: 1.088 (1.009,
1.173) per 1-SD (1.76 pg/dL)
increase of erythrocyte Pb
concentration
Males: 1.131 (1.027, 1.246)
per 1-SD (1.81 pg/dL)
increase of erythrocyte Pb
concentration
Females: 1.013 (0.886, 1.158)
per 1-SD (1.56 pg/dL)
increase of erythrocyte Pb
concentration
RR for Overall lymphoid
malignancy and erythrocyte
Pb quartiles:
Entire cohort:
Q1
Q2
Q3
Q4
Reference
1.35 (0.94, 1.95)
1.06 (0.71, 1.56)
1.52 (1.02, 2.25)
10-37
-------
Outcome
Confounders
Effect Estimates and 95%
Clsa
Males:
Q1:
Reference
Q2:
1.53 (0.93,
2.52)
Q3:
1.41 (0.84,
2.38)
Q4:
1.85 (1.10,
3.12)
Females:
Q1:
Reference
Q2:
0.98 (0.57,
1.67)
Q3:
1.04 (0.61,
1.78)
Q4:
0.92 (0.51,
1.65)
RR:
All B-cell NHL:
Entire cohort: 1.093 (1.005,
1.19) per 1-SD (1.76 pg/dL)
increase of erythrocyte Pb
concentration
Males: 1.151 (1.03, 1.286) per
1-SD (1.81 pg/dL) increase of
erythrocyte Pb concentration
Females: 1.013 (0.869, 1.18)
per 1-SD (1.56 pg/dL)
increase of erythrocyte Pb
concentration
Reference and
Study Design
Study Population
questionnaires
beginning in 1997.
Exposure Assessment
Males:
Q1: Oto <2.4736 pg/dL
Q2: 2.4736 to
<3.2876 pg/dL
Q3: 3.2876 to
<4.4026 pg/dL
Q4: >4.4026 pg/dL
Females:
Q1: Oto <1.8132 pg/dL
Q2: 1.8132 to
<2.5087 pg/dL
Q3: 2.5087 to
<3.6404 pg/dL
Q4: >3.6404 pg/dL
DLBCL
Entire cohort: 1.088 (0.943,
1.256) per 1-SD (1.76 pg/dL)
increase of erythrocyte Pb
concentration
Males: 1.07 (0.897, 1.276) per
1-SD (1.81 pg/dL) increase of
erythrocyte Pb concentration
Females: 1.183 (0.895, 1.565)
per 1-SD (1.56 pg/dL)
10-38
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
increase of erythrocyte Pb
concentration
CLL/SLL:
Entire cohort: 1.083 (0.916,
1.28) per 1-SD (1.76 pg/dL)
increase of erythrocyte Pb
concentration
Males: 1.274 (1.016, 1.598)
per 1-SD (1.81 pg/dL)
increase of erythrocyte Pb
concentration
Females: 0.736 (0.524, 1.034)
per 1-SD (1.56 pg/dL)
increase of erythrocyte Pb
concentration
FL:
Entire cohort: 1.397 (1.085,
1.798) per 1-SD (1.76 pg/dL)
increase of erythrocyte Pb
concentration
Males: 1.301 (0.951, 1.78) per
1-SD (1.81 pg/dL) increase of
erythrocyte Pb concentration
Females: 2.158 (1.07, 4.353)
per 1-SD (1.56 pg/dL)
increase of erythrocyte Pb
concentration
Other B-cell lymphoma:
Entire cohort:0.93 (0.717,
1.206) per 1-SD (1.76 pg/dL)
increase of erythrocyte Pb
concentration
Males: 1.022 (0.674, 1.549)
per 1-SD (1.81 pg/dL)
10-39
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Reference and
Study Design
Study Population
Exposure Assessment
Outcome
Confounders
Effect Estimates and 95%
Clsa
increase of erythrocyte Pb
concentration
Females: 0.803 (0.502, 1.284)
per 1-SD (1.56 pg/dL)
increase of erythrocyte Pb
concentration
MM:
Entire cohort: 1.114 (0.932,
1.332) per 1-SD (1.76 pg/dL)
increase of erythrocyte Pb
concentration
Males: 1.111 (0.887, 1.392)
per 1-SD (1.81 pg/dL)
increase of erythrocyte Pb
concentration
Females: 1.148 (0.81, 1.627)
per 1-SD (1.56 pg/dL)
increase of erythrocyte Pb
concentration
aEffect estimates are standardized to a 1 |jg/dL increase in blood Pb or a 10 |jg/g increase in bone Pb, unless otherwise noted. If the Pb biomarker is log-transformed, effect
estimates are standardized to the specified unit increase for the 10th-90th percentile interval of the biomarker level. Effect estimates are assumed to be linear within the evaluated
interval. Categorical effect estimates are not standardized.
bUnits assumed to be pg/dL (written as pg/L in the paper).
°Blood Pb measurements were converted from pg/Lto pg/dL.
dEffect estimates unable to be standardized.
fFrom 2013 Pb ISA.
ALAD = 6-aminolevulinic acid dehydratase; BLL = blood lead level; BMI = body mass index; CLL = Chronic Lymphatic Lymphoma; CLL/SLL = chronic lymphocytic leukemia/small
lymphocytic lymphoma; CPS-II = Cancer Prevention Study-ll (CPS-II) LifeLink Cohort; CRP = C-reactive protein; DLBCL = diffuse large B-cell lymphoma; EPIC- = European
Prospective Investigation into Cancer and Nutrition; FL = follicular lymphoma; GFAAS = graphite furnace atomic absorption spectrometry; GFR = glomerular filtration rate;
HR = hazard ratio; ICD = International Classification of Diseases; ICP-MS = inductively coupled plasma mass spectrometry; KNHANES = Korea National Health and Nutrition
Examination Survey; MM = multiple myeloma; NHANES = National Health and Nutrition Examination Survey; NHL = non-Hodgkin's lymphoma; NSDHS = Northern Sweden Health
and Disease Study; OR = odds ratio; Pb = lead; PIR = poverty-income ratio; RR = relative risk; SD = standard deviation; UC = urothelial carcinoma; WHO = World Health
Organization.
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EPA/600/R-23/375
APDA Environmental Protection Januaiy 2024
M m Agency www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 11: Effects of Lead in Terrestrial
and Aquatic Ecosystems
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
-------
CONTENTS
DOCUMENT GUIDE 11-iii
LIST OF TABLES 11 -v
LIST OF FIGURES 11-vi
ACRONYMS AND ABBREVIATIONS 11-vii
APPENDIX 11 EFFECTS OF LEAD IN TERRESTRIAL AND AQUATIC ECOSYSTEMS 11-1
11.1 Introduction, Scope, Concepts, and Tools 11-2
11.1.1 Scoping and Criteria for Study Inclusion 11-2
11.1.2 Introduction to Ecosystem Connections and Pb Transfers 11-6
11.1.3 Concentrations of Pb in Non-Air Media 11-7
11.1.4 Concepts Related to Ecosystem Effects of Pb 11-16
11.1.5 Ecosystem Services 11-17
11.1.6 Bioavailability 11-18
11.1.7 Risk Screening Tools 11-20
11.2 Terrestrial Ecosystems 11-24
11.2.1 Summary of New Information on Effects of Pb in Terrestrial Ecosystems and
Causality Determination Update Since the 2013 Pb ISA 11-24
11.2.2 Factors Affecting Bioavailability, Uptake and Bioaccumulation and Toxicity in
Terrestrial Biota 11-28
11.2.3 Environmental Concentrations of Pb in Terrestrial Biota and Ecosystems in the
United States at Different Locations and Over Time 11-51
11.2.4 Effects of Pb in Terrestrial Systems 11-54
11.2.5 Exposure and Response of Terrestrial Species 11-76
11.2.6 Terrestrial-Community and Ecosystem Effects 11-94
11.3 Freshwater Ecosystems 11-97
11.3.1 Summary of New Information on Effects of Pb in Freshwater Ecosystems and
Causality Determination Update Since the 2013 Pb ISA 11-97
11.3.2 Factors Affecting Bioavailability, Uptake and Bioaccumulation and Toxicity in
Freshwater Biota 11-103
11.3.3 Environmental Concentrations of Pb in Freshwater Biota and Ecosystems in the
United States at Different Locations and Over Time 11-120
11.3.4 Effects of Pb in Freshwater Systems 11-121
11.3.5 Exposure and Response of Freshwater Species 11-135
11.3.6 Freshwater-Community and Ecosystem Effects 11-161
11.4 Saltwater Ecosystems 11-164
11.4.1 Summary of New Information on Effects of Pb in Saltwater Ecosystems and Causality
Determination Update Since the 2013 Pb ISA 11-164
11.4.2 Factors Affecting Bioavailability, Uptake and Bioaccumulation, and Toxicity in
Saltwater Biota 11-168
11.4.3 Environmental Concentrations of Pb in Saltwater Biota in the United States at
Different Locations and Over Time 11-184
11.4.4 Effects of Pb in Saltwater Systems 11-186
11.4.5 Exposure and Response of Saltwater Species 11-192
11.4.6 Saltwater Community and Ecosystem Effects 11-205
11.5 References 11-208
11 -iv
-------
LIST OF TABLES
Table 11-1
Table 11-2
Table 11-3
Table 11-4
Pb concentration in non-air media and biota
11-8
Summary of Pb causality determinations for terrestrial plants, invertebrates, and
vertebrates 11-27
Studies of factors that affect the interpretability of exposure-response experiments in
terrestrial biota, since the 2013 Pb ISA 11-82
Summary of Pb causality determinations for freshwater plants, invertebrates, and
vertebrates
11-101
Table 11-5
Table 11-6
Table 11-7
Studies in freshwater biota with analytically verified Pb concentrations and that report an
effect on growth, reproduction or survival comparable to, or lower than, the lowest effect
concentrations reported in previous Pb AQCDs or the 2013 Pb ISA 11-142
Updated causality determinations for Pb in saltwater organisms and ecosystems_
11-167
Studies in saltwater biota with analytically verified Pb concentration that report an effect on
growth, reproduction, or survival comparable to, or lower than, the lowest effect
concentrations reported in previous Pb AQCDs or the 2013 Pb ISA 11-198
11-v
-------
LIST OF FIGURES
Figure 11 -1 Locations of the 4,857 soil sampling sites included in the U.S. Geological Survey North
American Soil Geochemical Landscapes Project conducted from 2007 to 2010. 11-13
Figure 11 -2 Conceptual diagram for evaluating bioavailability processes and bioaccessibility for metals
in soil, sediment, or aquatic systems. 11-19
Figure 11-3 Change in toxicity expressed as relative responses (i.e., response relative to the mean of
the corresponding control soil) for three different laboratory soil treatments: freshly spiked;
spiked, leached and pH-corrected; and spiked, leached and pH-corrected with 5 years of
aging. 11-33
Figure 11 -4 Maps of Pb sampled from A-horizon (A) and C-horizon (B) soils, the ratio of Pb observed
in A-horizon to C-horizon soils (C) and a map of U.S. population density (D). 11-53
Figure 11-5 Main forms of Pb in seawater as a function of pH at 25°C and salinity of 35 ppt. 11-171
Figure 11-6 Acute genus sensitivity distribution for saltwater biota from Church et al. (2017). 11-196
Figure 11 -7 Comparison of chronic sensitivity distributions in saltwater biota for dissolved Pb following
the U.S. EPA and European Union methods. 11-197
11 -vi
-------
ACRONYMS AND ABBREVIATIONS
ACE abundance-based coverage estimator
AChE acetylcholinesterase
Ag silver
ALAD aminolevulinic acid dehydratase
AMF arbuscular mycorrhizal fungi
AQCD Air Quality Criteria Document
As arsenic
ASTM American Society for Testing and
Materials
AVS acid volatile sulfide
AWCD average cell wall color development
AWQC ambient water quality criteria
BAF bioaccumulation factor
BCF bioconcentration factor
BEST Biomonitoring of Environmental Status
and Trends
BLL blood lead level
BLM biotic ligand model
BMF biomagnification factors
BRT boosted regression tree
BSAF biota-sediment accumulation factor
Ca calcium
CAT catalase
CCA canonical correspondence analysis
CCC criterion continuous concentration
Cd cadmium
CEC cation exchange capacity
CF conversion factor
CMC criteria maximum concentration
CORT corticosterone
CRADA Cooperative Research and
Development Agreement
CSMW California State Mussel Watch
Cu copper
d day(s)
DOC dissolved organic carbon
dpf days postfertilization
dph days posthatch
DOM dissolved organic matter
DT diatom + tetramin
EC50 half maximal effect concentration
eCEC effective cation exchange capacity
Eco-SSL ecological soil screening level
EDTA ethylenediaminetetraacetic acid
FCORT fecal corticosterone
FCV final chronic value
GABA gamma-aminobutyric acid
GPx glutathione peroxidase
GSE1 glutathione
GST glutathione-S-transferase
ElAB harmful algal bloom
hpf hours postfertilization
IC inhibitory concentration
ISA Integrated Science Assessment
Kd partition coefficient
LECES Level of Biological Organization,
Exposure, Comparison, Endpoint, and
Study Design
LH luteinizing hormone
LOEC lowest observed effect concentration
LOAEL lowest observed adverse effect level
LRMN Large River Monitoring Network
LUFA Landwirtschaftliche Untersuchungs-
und Forschungsanstalt
MATC maximum acceptable toxicant
concentration
MBC microbial biomass carbon
MDA malondialdehyde
ME mining ecotype
Mg magnesium
MIC minimum inhibitory concentration
MLR multiple linear regression
mo month(s)
MRG metal-rich granules
MTC maximum tolerable concentration
MW molecular weight
NAAQS National Ambient Air Quality
Standards
NASGLP North American Soil Geochemical
Landscapes Project
NAWQA National Water Quality Assessment
NEC no-effect concentration
NME nonmining ecotype
NOAA National Oceanic and Atmospheric
Administration
NOEC no-observed-effect concentration
NOM natural organic matter
n.s. nonsignificant
OC organic carbon
OM organic matter
OP omnivores-predator
11-vii
-------
OTU operational taxonomic unit TBMF
Pb lead TEC
PEC probable effects concentration TRF
PECOS Population, Exposure, Comparison, TTF
Outcome and Study Design jj g
PMF Picher mine field
PNEC predicted no-effect concentration USGS
REACH Registration, Evaluation, Authorisation WACAP
and Restriction of Chemicals
ROS reactive oxygen species WEOC
SEM simultaneously extracted metal wk
SOD superoxide dismutase WQC
SSD species sensitivity distribution YCT
T3 triiodothyronine yr
T4 thyroxine Zn
trophic biomagnification factor
threshold effect concentration
terminal restriction fragment
trophic transfer factor
United States Environmental Protection
Agency
United States Geological Survey
Western Airborne Contaminants
Assessment Project
water-extractable organic carbon
week(s)
water quality criteria
yeast, cereal leaves, and trout pellet
year(s)
zinc
11-viii
-------
APPENDIX 11 EFFECTS OF LEAD IN TERRESTRIAL AND
AQUATIC ECOSYSTEMS
Summary of Causality Determinations for Welfare Effects of Lead
This appendix characterizes the scientific evidence that supports causality determinations for lead
(Pb) exposure and the effects of Pb in terrestrial and aquatic ecosystems and biota. In assessing the overall
evidence, the strengths and limitations of individual studies were evaluated. More details on the causal
framework used to reach these conclusions are included in the Preamble to the Integrated Science
Assessments (U.S. EPA. 2015). The evidence presented throughout this appendix supports the following
causality determinations (bolded text indicates a change since the 2013 Integrated Science Assessment for
Pb).
Level
Effect
Terrestrial3
Freshwater3
Saltwater3
Community-
and Ecosystem
Community and Ecosystem Effects
Likely Causal
Likely Causal
Suggestive
Reproductive and Developmental Effects - Plants
Inadequate
Inadequate
Inadequate
(0
+-»
C
Reproductive and Developmental Effects -
Invertebrates
Causal
Causal
Likely
Causal
O
Q.
¦o
C
LD
<1)
C
o
Reproductive and Developmental Effects -
Vertebrates
Causal
Causal
Inadequate
0
>
Growth - Plants
Causal
Likely Causal
Inadequate
_l
1
C
Growth - Invertebrates
Likely Causal
Causal
Inadequate
O
+-»
03
1 8>
Hematological Effects - Vertebrates
Causal
Causal
Inadequate
c c
03 O
O) Q.
Physiological Stress - Plants
Causal
Likely Causal
Inadequate
5 8
= *
Physiological Stress - Invertebrates
Likely Causal
Likely Causal
Suggestive
w
Physiological Stress - Vertebrates
Likely Causal
Likely Causal
Inadequate
aBased on the weight of evidence for causal determination in Table II of the ISA Preamble (U.S. EPA. 2015).
The Executive Summary, Integrated Synthesis, and all other appendices of this Pb ISA can be found at
https://assessments.epa.gov/isa/document/&deid=359536.
11-1
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11.1 Introduction, Scope, Concepts, and Tools
This appendix synthesizes and evaluates the most policy-relevant scientific information on Pb
welfare effects to help form the foundation for the review of the secondary (welfare1-based) National
Ambient Air Quality Standards (NAAQS) for lead (Pb). The focus of this appendix is on studies
published since the 2013 Integrated Science Assessment (ISA) for Pb (2013 Pb ISA) U.S. EPA (2013)
that examine Pb interactions with the biotic components of terrestrial and aquatic ecosystems, including
effects on vegetation and wildlife. Pb transport through abiotic compartments (air, soil, water, and
sediment) is covered in Appendix 1: Lead Source to Concentration:
https://asscssmcnts.cpa.go\/isa/documcnt/&dcid=359536. Section 11.1 of this appendix includes key
concepts and tools useful for characterizing the effects of Pb on biota. Section 11.2 examines the
bioavailability, bioaccumulation, and effects of Pb in terrestrial ecosystems. The effects of Pb in
terrestrial environments are followed by information on the bioavailability, bioaccumulation and effects
of Pb in freshwater (Section 11.3) and saltwater (Section 11.4) ecosystems.
11.1.1 Scoping and Criteria for Study Inclusion
This appendix builds upon the assessment of effects of Pb on ecosystems reported in the 2013 Pb
ISA (U.S. EPA, 2013) and in prior Air Quality Criteria Documents (AQCDs) from 1977 (U.S. EPA,
1977), 1986 (U.S. EPA, 1986), and 2006 (U.S. EPA, 2006). The framework used to define the scope of
the ecological effects portion of the current ISA is modeled after the Population, Exposure, Comparison,
Outcome, and Study Design (PECOS) used for human health effects (Appendix 12:
https://asscssmcnts.cpa.go\/isa/documcnt/&dcid=3 59536). For the health appendices, the PECOS
statement defines the objectives of the review and establishes study inclusion criteria, thereby facilitating
identification of the most relevant literature to inform the ISA for each health discipline. Similarly, the
Level of Biological Organization, Exposure, Comparison, Endpoint, and Study Design (LECES)
statement aids in identifying the relevant evidence in the literature for the ecological effects of Pb
(Table 12-4; Appendix 12). Studies that reported the effects of Pb on biota were evaluated, included, and
discussed in this appendix if they satisfied the following LECES criteria:
11.1.1.1 Level of Biological Organization
Studies considered for this appendix included those that reported Pb effects on species,
subspecies or populations of vegetation, microbes, invertebrates, or vertebrates at any lifestage on
'Under The Clean Air Act (CAA) section 302(h) (42 U.S.C. § 7602(h)), effects on welfare include, but are not
limited to, "effects on soils, water, crops, vegetation, manmade materials, animals, wildlife, weather, visibility, and
climate, damage to and deterioration of property, and hazards to transportation, as well as effects on economic
values and on personal comfort and well-being."
11-2
-------
biological communities or on ecosystems in terrestrial, freshwater, or saltwater environments and
transition zones present in the United States or similar to those in the United States. In the 2013 Pb ISA,
ecological effects were generally organized in order of increasing biological complexity (i.e., from the
subcellular and cellular levels through the individual organism and up to ecosystem-level effects) (U.S.
EPA. 2013). This appendix follows the same organizing principle. For effects that occur at the
suborganism scale such as perturbation of biomarkers of physiological stress or changes in hematological
parameters, emphasis was placed on studies that concurrently reported effects experimentally linked to
higher levels of biological organization. Organism-level endpoints such as growth, survival, and
reproductive output have been definitively linked to effects at the population level and above. Examples
of organism-level endpoints with direct links to population-level effects include mortality, gross
abnormalities, survival, fecundity, and growth (Suter et al.. 2004). Because of the complexity of processes
that can affect an ecosystem and considering that Pb rarely occurs as the only contaminant in natural
systems, it is difficult to attribute effects observed at higher levels of biological organization solely to Pb.
11.1.1.2 Exposure
The deleterious effects of any given concentration of Pb can vary greatly under different
environmental and experimental conditions, as well as the duration and pathway of exposure. Relevant
concentrations for this assessment take into consideration the range of Pb concentrations in environmental
media from U.S. locations (Table 11-1) and the available evidence for concentrations at which effects are
observed in microbes, plants, invertebrates, and vertebrates. Effects observed at or near environmental
concentrations of Pb measured in soil, sediment, and water are emphasized. For the studies included in
the 2013 Pb ISA, evidence from exposures or doses generally ranged "from current levels to one or two
orders of magnitude above current levels" (U.S. EPA. 2013). Statements regarding concentrations
considered for literature inclusion were not provided in earlier United States Environmental Protection
Agency (U.S. EPA) reviews of this metal. To focus on studies that are the most policy-relevant with
regard to current environmental concentrations of Pb in the United States, concentration guidelines were
applied when evaluating the literature published since the 2013 Pb ISA (Appendix 12:
https://asscssmcnts.cpa.go\/isa/documcnt/&dcid=359536). These guidelines were derived by taking into
consideration data that was current at the time of the 2013 Pb ISA on Pb concentrations in soils, water and
sediments in the United States. The concentration guideline for literature screening in this ISA is
approximately one order of magnitude higher than upper bound values from available environmental
surveys (Table 1-1 from 2013 Pb ISA). For soil, the concentration guideline for screening of terrestrial
studies of Pb exposure and effects was set at approximately 230 mg Pb/kg of soil, although higher
concentrations were considered if the study added new information on a mechanism of action, or if the
higher concentration was part of a series that contributed exposure-response information and included
other concentrations below 230 mg Pb/kg. The concentration guideline for screening aqueous exposure
studies was approximately 10 |Lig Pb/L, although higher concentrations were considered if the study added
new information on a mechanism of action or if the higher concentration was part of a series that
11-3
-------
contributed exposure-response information. The concentration guideline for screening sediment studies
was approximately 300 mg Pb/kg dry weight or lower. Studies at very high concentrations of Pb were
excluded unless they were part of a series in an experimental exposure-response study and at least one
concentration in the test series was in the ranges stated above. The approach for selection of the Pb
concentrations used as guidelines and additional information on scoping for the literature for ecological
effects of Pb is provided in Appendix 12. Initial literature search and screening steps for this review
identified many studies conducted at higher concentrations of Pb that were ultimately excluded from the
draft ISA (https://hero.epa.gov/hero/index.cfm/proiect/page/proiect id/4081).
For references published since the 2013 Pb ISA, values reported in this appendix for biological
effects are from exposures in which concentrations of Pb were analytically verified (measured). Some
nominal concentrations are cited but the studies where they are used are experimental gradient studies
where the response of an organism or system to a series of increasing exposures is informative with
respect to causality. In no case are such studies cited to support quantification of effect concentrations or
quantification of the exposure-response relationship. In addition, many studies in terrestrial and aquatic
environments report a series of nominal additions of Pb to the environmental medium, but measure the
concentrations present in the organisms and analyze their effects on organisms, not the relationship
between the medium concentration and effects on organisms. Older studies cited in prior AQCDs or in the
2013 Pb ISA and occasionally referenced in this appendix may have used nominal exposures. For
consistency, concentrations of Pb in soil and sediment are reported in mg Pb/kg dry weight (unless
otherwise specified) and aqueous concentrations of Pb are reported as ng Pb/L. For study concentrations
originally in other units such as (iM or ppb, the values are converted to mg Pb/kg or |ig Pb/L, and original
reported units are retained in parentheses. Only a subset of the studies reporting Pb effects on biota
analytically verified the concentration of Pb in media and the test organisms investigated.
11.1.1.3 Comparison
Comparisons in the studies considered for inclusion in this appendix were to an unexposed laboratory
control, a reference population, or a site with no detectable exposure or with lower Pb exposure. For
ecological effects assessment, both laboratory and field studies (including field experiments and
observational studies) can provide useful data (U.S. EPA. 2015). As the number of factors that the study
holds constant increases, other than Pb exposure, so does the certainty with which observed variation in
outcomes can be attributed to exposure, while the size of effects that the study is capable of attributing to
exposure becomes smaller. The ability to hold other variables constant is expected to diminish with
increasing biological scale from subcellular processes to whole ecosystems and from laboratory to field.
In general, effects of Pb on ecological endpoints are reported in the ISA if they are statistically
significant.
11-4
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11.1.1.4 Endpoint
The biological endpoints considered in this appendix are relevant to the levels of biological
organization discussed above. The endpoints encompass individual organism-level or population-level
effects on a given species including but not limited to effects on growth, reproduction or development,
neurobehavioral effects, reduced survival, or fitness, and photosynthesis. At higher levels of biological
organization, endpoints include, but are not limited to, changes in community composition (e.g., shifts in
genotypes or species, species extirpation, declines in the total number of species, and decreased species
richness), declines in biomass, and other altered ecosystem processes and functions.
11.1.1.5 Study Design
Relevant study designs for assessing Pb effects on ecological receptors include laboratory,
mesocosm, observational or experimental field or gradient studies wherein observed effects are measured
and analyzed quantitatively, or mechanistic modeling studies that estimate the effect of Pb on an
organism, biological population, community, or ecosystem (U.S. EPA, 2015). Controlled exposure
studies in laboratory or small-to-medium-scale field settings provide the most direct evidence for the
effects of Pb exposure, but their scope of inference may be limited (U.S. EPA, 2013). Exposure-response
data from acute bioassays typically report effects on mortality, growth, or reproduction. Chronic
bioassays are designed to incorporate effects over the lifespan or partial lifespan of the study subjects,
including effects on reproduction. In contrast, mesocosms and field studies include potentially
confounding factors (e.g., other metals) or factors known to interact with exposure (e.g., pH), thus
increasing the uncertainty in associating the effects observed with exposure to Pb (U.S. EPA, 2013).
11.1.1.6 Additional Scoping
Topics within scope also include effects of Pb biogeochemistry on bioavailability in terrestrial,
freshwater, and saltwater environments as well as subsequent vulnerability of particular organisms,
populations, communities or ecosystems and studies that address key uncertainties and limitations in the
evidence identified in the previous review. Topics outside of the scope of this appendix included mixture
studies that did not assess Pb effects independently and site-specific studies in non-U.S. locations that do
not contribute novel insights on Pb biogeochemistry or effects. As in the 2013 Pb ISA, generally, studies
on mine tailings, industrial effluent, sewage, bioremediation of highly contaminated sites and ingestion of
Pb shot, pellets or fishing gear are not within the scope of this ISA due to the high concentration of Pb
and lack of a connection to air-related sources or processes. This is consistent with the 2006 AQCD,
which typically did not include "effects from irrelevant exposure conditions relative to airborne emissions
of Pb (e.g., Pb shot, Pb paint, injection studies, studies conducted on mine tailings and studies conducted
with hydroponic solutions)" (Section AX 7.1.3 of (U.S. EPA. 2006)).
11-5
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11.1.2
Introduction to Ecosystem Connections and Pb Transfers
Metals, including Pb, occur naturally in the geosphere, and anthropogenic enrichment of these
elements can lead to elevated concentrations in terrestrial and aquatic ecosystems. Pb is a persistent metal
that, once emitted, may cycle through multiple environmental media compartments (e.g., air, soil, water,
sediment) prior to exposure to plants and animals, as discussed in Appendix 1: https://assessments.epa.
gov/isa/document/&deid=359536 (Section 1.3). In terrestrial ecosystems, non-air media can receive Pb
from atmospheric deposition or other sources. The contribution of atmospheric Pb differs by location and
there is a lack of source apportionment studies to characterize the amount of Pb from deposition in
relation to other Pb sources. Once deposited, Pb can be resuspended into the air or transferred among
other environmental media (Section 1.3). Exposure of freshwater and estuarine organisms to Pb, and
associated effects, are tied to terrestrial systems via watershed processes. Atmospherically derived Pb can
enter aquatic systems through erosional transport of soil particles in runoff from terrestrial systems
(Section 1.3.3) or via direct wet or dry deposition over a water surface (Section 1.3.1.2). Once in the
aquatic environment, Pb partitions between various compartments (water column, sediment, biota;
Section 1.3.3). Saltwater ecosystems include habitats that encompass a range of salinities from just above
that of freshwater to that of seawater. These ecosystems may receive Pb contributions from atmospheric
deposition (Section 1.3.1.2). riverine transport (Section 1.3.3) and runoff (Section 1.3.3) from terrestrial
systems. Ecosystems in more urban areas are also influenced by non-air sources of Pb such as paint,
automobiles, wastewater, and industrial activities. Although Pb is present in the natural environment, it
has no biological function in plants or animals. Terrestrial, freshwater, and marine/estuarine organisms
have developed adaptive physiological responses for living with metals. These adaptations may include
intracellular sequestration (e.g., synthesis of metallothioneins or metal-rich granules [MRG]), induction of
enzymes involved in oxidative stress response, and modification of metal uptake or elimination rates
(Gismondi et al.. 2017). Anthropogenic enrichment can result in concentrations that exceed the capacity
of organisms to regulate internal concentrations, causing a toxic response and potentially death. Across
taxa, effects of Pb exposure are likely mediated through common biological mechanisms. In the case of
Pb, ecological receptors and humans are linked via shared pathways of exposure and commonalities in
biological response to this metal (Lassiter et al.. 2015).
Connections between the atmosphere, the abiotic and biotic compartments of terrestrial and
aquatic ecosystems, and humans are acknowledged for Pb. However, for the purposes of this ISA, these
topics are divided into different appendices. Within this Ecological Effects appendix, terrestrial,
freshwater, and saltwater ecosystems are considered separately because of different environmental and
physiological factors that influence Pb toxicity, such as bioavailability of the metal, form of Pb, other
water and soil chemistry factors, and organism adaptations.
11-6
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11.1.3
Concentrations of Pb in Non-Air Media
Organisms may be exposed to Pb in soil, water, sediment, and other biota (via diet). Food,
drinking water, and contact with contaminated soils are likely major routes of exposure for terrestrial
wildlife. Ingestion and water intake are major routes of exposure for aquatic fauna. Inhalation is thought
to be a minor pathway in wildlife, with the possible exception of exposures in proximity to Pb
atmospheric point sources, such as smelters. Due to the presence of Pb in various environmental media,
exposure to this metal can occur via multiple pathways.
To provide sufficient information to support development of air quality criteria for Pb that are
protective of terrestrial and aquatic systems, it is important to gain a general understanding of current
distribution and the concentrations of Pb in the environment. Information on environmental
concentrations of Pb at U.S. locations is tabulated in Table 11-1. This table updates Table 6-2 in the 2013
Pb ISA U.S. EPA (2013) on Pb concentration in non-air media and biota. Sources of environmental
concentration data in Table 11-1 were limited to regional or national-scale studies. Studies that reported
concentrations in environmental media for one or a very small number of locations would be considered
anecdotal for the purpose of this review. Measured concentrations of Pb in soils, sediment and water are
not necessarily representative of the amount of Pb available to elicit a toxic effect. For Pb to interact with
a biological membrane and be taken up into an organism, it must be in a bioavailable form
(Section 11.1.6), which is dependent upon the physical, chemical, and biological conditions under which
an organism is exposed at a particular geographic location. In addition, caution must be taken while
comparing Pb concentrations in different studies of environmental media because reported concentrations
of Pb may not be directly comparable across studies, in part due to differences in sampling, collection and
measurement methods. For example, soil Pb measurements may vary between studies that used partial
and complete acid digestion. Furthermore, complete acid digestion is likely to overestimate the amount of
bioavailable Pb in many cases. In aquatic systems, measurements of dissolved Pb may vary among
collection methods, notably due to different sample filtration sizes, while the composition of sediment
samples of Pb is often influenced by sieving size. These are given as illustrative examples of how Pb
observations may be affected by methods, but a comprehensive discussion of Pb sampling, collecting, and
measuring methods is beyond the scope of this ISA.
Some surveys of Pb in environmental media in Table 11-1 predate the 2013 Pb ISA (U.S. EPA,
2013) and 2006 Pb AQCD (U.S. EPA, 2006). Although they may have used less optimal methods than
more recent studies, these data are not excluded from the ISA in cases wherein they remain the best
available information.
11-7
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Table 11-1
Pb concentration in non-air media and biota
Media
Pb Concentration
Years Data
Obtained
References
Soil
Conterminous U.S. 0-5 cm depth soil:
Median: 18.1 ± 185 mg Pb/kg; range: <0.5-
12,400 mg Pb/kg; IQR: 13.5-23.9 mg Pb/kg (dry weight)
Conterminous U.S. A horizon soil:
Median: 17.8 ± 46.6 mg Pb/kg; range: <0.5-
2,200 mg Pb/kg; IQR: 13.2-23.2 mg Pb/kg (dry weight)
Conterminous U.S. C horizon soil:
Median: 14.9 ± 18.5 mg Pb/kg; range: <0.5-681 mg Pb/kg;
IQR: 11.1-19.2 mg Pb/kg (dry weight)
2007-2010
Smith et al.
(2013a)
Northeastern U.S. forest floor soil mean: 151 ±29 mg Pb/kg
(dry weight)
1980
Richardson et
al. (2014)
Northeastern U.S. forest floor soil mean: 68 ± 13 mg Pb/kg
(dry weight) (resurvey of 16 of 25 1980 sites)
2011
Richardson et
al. (2014)
Soil sampled at 54 sites in Los Angeles, Orange, San
Bernardino, and Riverside counties in California
Range: 5-70 mg Pb/kg
Mean: 23.9 ± 13.8 mg Pb/kg
2019
Mackowiak et
al. (2021)
Soil (freshwater
wetlands and salt
marshes)
Conterminous U.S. uppermost soil horizon mean:
20.15 ± 1.73 (95% CI) mg Pb/kg (dry weight)
2011
Nahlik et al.
(2019)
Cores from 35 U.S. lakes
1970s Median: 115 mg Pb/kg (dry weight)
1990s Median: 73 mg Pb/kg (dry weight)
1996-2001
Mahler et al.
(2006)
National Water Quality Assessment of lotic systems
Median: 28 mg Pb/kg (dry weight)
1991-2003
U.S. EPA
(2006)
Freshwater
Sediment
National Water Quality Assessment of lotic systems
grouped by river basin land use:
Baseline (in low-population areas): median: 20 mg Pb/kg;
range: 2-200 mg Pb/kg (dry weight)
Agricultural sites: median: 20 mg Pb/kg; range: 6-
310 mg Pb/kg (dry weight)
Cropland sites: median: 19 mg Pb/kg; range: 8-
310 mg Pb/kg (dry weight)
Pasture sites: median: 20 mg Pb/kg; range: 6-49 mg Pb/kg
(dry weight)
Forested sites: median: 28 mg Pb/kg; range: 2-
200 mg Pb/kg (dry weight)
Rangeland sites: median: 18 mg Pb/kg; range: 6-
330 mg Pb/kg (dry weight)
1991-2001
Horowitz and
Stephens
(2008)
131 coastal conterminous U.S. rivers: Overall mean:
59 mg Pb/kg; median: 26 mg Pb/kg (dry weight)
Atlantic rivers: mean: 110 mg Pb/kg; median: 36 mg Pb/kg
(dry weight)
Gulf rivers: mean: 32 mg Pb/kg; median: 24 mg Pb/kg (dry
weight)
2010-2011
Horowitz et al.
(2012)
11-8
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Media Pb Concentration
Years Data
Obtained
References
Pacific rivers: mean: 19 mg Pb/kg; median: 13 mg Pb/kg
(dry weight)
Global Range: 0.6-1,050 mg Pb/kg
U.S. Range (from Puget Sound): 13.4-52.8 mg Pb/kg
Reported in
studies dated
1977-1990
Sadia (1992)
Saltwater Sediment U.S. Geometric Mean: 43 mg Pb/kg
Global Geometric Mean: 43 mg Pb/kg
Global Geometric Mean ("hot spot" data from contaminated
sites removed): 34 mg Pb/kg
1984-1987
Cantillo and
O'Connor
(1992)
Median: 0.50 pg Pb/L
Max: 30 pg Pb/L, 95th percentile 1.1 pg Pb/L
1991-2003
U.S. EPA
(2006)
Fresh Surface
Water
(Dissolved Pb)
8 Texas rivers
Sabine: Mean: 0.04 ± 0.025 |jg Pb/L
Range: 0.013-0.098 |jg Pb/L
Neches: Mean: 0.036 ± 0.028 |jg Pb/L
Range: 0.01-0.099 |jg Pb/L
Trinity: Mean: 0.061 ± 0.067 |jg Pb/L
Range: 0.009-0.218 |jg Pb/L
Brazos: Mean: 0.02 ± 0.011 |jg Pb/L
Range: 0.008-0.061 |jg Pb/L
Colorado: Mean: 0.02 ± 0.009 |jg Pb/L
Range: 0.007-0.04 |jg Pb/L
Guadalupe: Mean: 0.049 ± 0.059 |jg Pb/L
Range: 0.005-0.202 |jg Pb/L
San Antonio: Mean: 0.356 ± 0.235 |jg Pb/L
Range: 0.177-0.919 pg Pb/L
Nueces/Frio: Mean: 0.025 ± 0.034 pg Pb/L
Range: 0.008-0.166 pg Pb/L
1997-1998
Jiann et al.
(2013)
Range: 0.0003-0.075 pg Pb/L
(Set of National Parks in western U.S.
Field and
2002-2007 Sherrell (2003)
Blett (2010)
Appalachian headwater streams
(4 sites located in second- or third-order streams within the
Blue Ridge level III ecoregion)
Mean: <0.28 pg Pb/L
2015-2017
Olson et al.
(2019)
8 Texas rivers
Sabine: Mean: 27.76 ± 5.5 mg Pb/L
Range: 21.81-38.17 mg Pb/L
Neches: Mean: 32.4 ± 4.55 mg Pb/L
Fresh Surface
Water (Particulate
Pb)
Range: 26.48-39.23 mg Pb/L
Trinity: Mean: 28.24 ± 3.82 mg Pb/L
1 gg7 1998 J'ann e* a^-
(2013)
Range: 22.87-33.24 mg Pb/L
Brazos: Mean: 22.45 ± 7.39 mg Pb/L
Range: 12.18-40.06 mg Pb/L
Colorado: Mean: 25.39 ± 12.33 mg Pb/L
11-9
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Media
Pb Concentration
Years Data
Obtained
References
Range: 13.4-72.92 mg Pb/L
Guadalupe: Mean: 20.2 ± 5.17 mg Pb/L
Range: 14.2-35.8 mg Pb/L
San Antonio: Mean: 28.8 ± 5.23 mg Pb/L
Range: 21.97-38.34 mg Pb/L
Nueces/Frio: Mean: 22.33 ± 4.67 mg Pb/L
Range: 14.05-32.27 mg Pb/L
Saltwater
Global Range: 0.01-27 |jg Pb/L
Open-Ocean Range: 0.01-4.8 |jg Pb/L
Reported in
studies dated Sadiq (1992)
1977-1990
Vegetation
Lichens: 0.3-5 mg Pb/kg (dry weight) (Set of National Parks
in western U.S.)
2002-2007 Blett (2010)
Leaves from woody shrubs and trees from 54 sites in Los
Angeles, Orange, San Bernardino and Riverside counties in
California
Adenostoma fasciculatum
Mean: 0.17 ± 0.08 (SE) mg Pb/kg
Artemisia caiifornica
Mean: 0.16 ± 0.01 (SE) mg Pb/kg
Baccharis salicifolia
Mean: 0.22 ± 0.03 (SE) mg Pb/kg
Encelia farinosa
Mean: 0.20 ± 0.02 (SE) mg Pb/kg
Eriogonum spp.
Mean: 0.23 ± 0.03 (SE) mg Pb/kg
Heteromeles arbutifolia
Mean: 0.42 ± 0.17 (SE) mg Pb/kg
Malosma luarina
Mean: 0.38 ± 0.06 (SE) mg Pb/kg
Quercus agrifolia
Mean: 0.29 ± 0.04 (SE) mg Pb/kg
2019
Mackowiak et
al. (2021)
Vertebrates
Fish (sampled from 111 sites in 9 river basins of large U.S.
rivers):
Mean: 0.07 mg Pb/kg (wet weight) (whole fish); Median:
0.10 mg Pb/kg (wet weight) (whole fish); 85th percentile:
0.27 mg Pb/kg (wet weight) (whole fish); Max:
9.29 mg Pb/kg (wet weight) (whole fish)
1995-2004
Hinck et al.
(2009)
Fish (96 sites in large U.S. rivers):
Female bass (Micropterus spp.): median: 0.04 mg Pb/kg;
mean: 0.06 ± 0.02 mg Pb/kg (wet weight) (whole fish)
Male bass (Micropterus spp.): median: 0.03 mg Pb/kg;
mean: 0.05 ± 0.01 mg Pb/kg (wet weight) (whole fish)
Female carp (Cyprinus carpio)'. median: 0.10 mg Pb/kg;
mean: 0.11 ± 0.01 mg Pb/kg (wet weight) (whole fish)
Male carp (Cyprinus carpio)'. median: 0.09 mg Pb/kg; mean:
0.12 ± 0.01 mg Pb/kg (wet weight) (whole fish)
1995-2004
Hinck et al.
(2008)
11-10
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Media
Pb Concentration
Years Data
Obtained
References
Dolphinfish (Coryphaena hippurus) in southern Gulf of
California (wet weight) (muscle tissue):
Mean: 0.059 mg Pb/kg
onnfi omr Gil-Manrique et
2006"2015 al. (2022)
Fish (from a set of national parks in western U.S.):
0.0033 (fillet) to 0.97 (liver) mg Pb/kg (dry weight)
2002-2007 Blett (2010)
Anna's hummingbirds (Calypte anna) surveyed in coastal,
valley and Sierra Nevada foothills regions of northern
California
Mean: 0.23 ± 0.25 mg Pb/kg; range: 0.00-1.35 mg Pb/kg
(body feathers; live) (dry weight)
Mean: 3.00 ± 7.64 mg Pb/kg; range: 0.28-46.0 mg Pb/kg
(body feathers; carcasses) (dry weight)
Mean: 1.01 ± 3.10 mg Pb/kg; range: 0.01-16.9 mg Pb/kg
(liver) (dry weight)
Mean: 0.94 ± 2.07 mg Pb/kg; range: 0.03-12.43 mg Pb/kg
(kidney) (dry weight)
Mean: 8.17 ± 36.27 mg Pb/kg (combined feathers) (dry
weight)
2015
Mikoni et al.
(2017)
Neotropic Cormorants (Phalacrocorax brasilianus) surveyed
in Lake Livingston, Texas:
Female mean: 4.92 ±4.11 (SE) mg Pb/kg (breast feathers)
(dry weight)
Male mean: 1.68 ± 0.822 (SE) mg Pb/kg (breast feathers)
(dry weight)
In Richland Creek Wildlife Management Area, Texas:
Female mean: 0.191 ± 0.044 (SE) mg Pb/kg (breast
feathers) (dry weight)
Male mean: 0.115 ± 0.015 (SE) mg Pb/kg (breast feathers)
(dry weight)
2014
Mora et al.
(2021)
Invertebrates
7 earthworm species in northeastern U.S.
Overall mean: 29 ± 6 (SE) mg Pb/kg (dry weight)
Amynthas agrestis mean: 21 ±11 (SE) mg Pb/kg (dry
weight)
Aporrectodea rosea mean: 43 ± 5 (SE) mg Pb/kg (dry
weight)
Aporrectodea tuberculata mean: 30 ± 7 (SE) mg Pb/kg (dry
weight)
Dendrobaena octaedra mean: 43 ± 20 (SE) mg Pb/kg (dry
weight)
Lumbricus rubellus mean: 24 ± 5 (SE) mg Pb/kg (dry
weight)
Lumbricus terrestris mean: 14 ± 4 (SE) mg Pb/kg (dry
weight)
Octolasion cyaneum mean: 20 ± 8 (SE) mg Pb/kg (dry
weight)
2013
Richardson et
al. (2015b)
Oysters (Crassostrea virginica) and mussels (Mytilus edulis)
in east coast U.S.
Range: 0.11-2.2 mg Pb/kg Pb (dry weight)
2003-2006
Shiel et al.
(2012)
11-11
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Media
Pb Concentration
Years Data
Obtained
References
Oysters (Crassostrea gigas) in west coast Canada
Range: 0.05-0.22 mg Pb/kg Pb (dry weight)
2002-2004
Shiel et al.
(2012)
CI = confidence interval; IQR = Interquartile range; Pb = lead; SE = Standard error.
This table updates Pb non-air media and biota concentration data from Tables 1-1 and 6-2 in the 2013 Pb ISA (U.S. EPA. 2013).
Sources of concentration data are limited to regional or national-scale studies.
Several large-scale surveys of soil Pb concentrations were identified for inclusion in the ISA. The
United States Geological Survey (USGS) North American Soil Geochemical Landscapes Project
(NASGLP) (Smith et al.. 2013a) is a recent soil survey that supplants Shacklette and Boerngen (1984).
the national soil survey cited in the 2013 Pb ISA, because of the larger size and extent, use of modern
geostatistical sampling methods, increased sampling resolution and documented data quality validation
(Smith et al.. 2013a). Shacklette and Boerngen (1984) collected 1,319 samples of Pb at a depth of 20 cm
along U.S. roadways between 1961 and 1976. The NASGLP provides a more comprehensive survey of
soil Pb in the conterminous United States because the survey employed a spatially balanced, sampling-
location selection method and collected soil samples from multiple depths at each selected location.
Samples were taken from depths of 0-5 cm in A-horizon and C-horizon soils at 4,857 sites systematically
selected using a generalized random tessellation stratified design in 2007-2010 (Figure 11-1). Soil Pb
concentrations were determined by inductively coupled plasma atomic emission spectroscopy and
inductively coupled plasma mass spectrometry analyses. Measurements were validated using documented
quality assurance and quality control procedures. A review of seven national-scale geochemical datasets
compared the NASGLP survey design to that of Shacklette and Boerngen (1984) and discussed the
methodological issues with other prior national-scale geochemical surveys that NASGLP was designed to
address (Smith et al.. 2013b). Summary statistics of conterminous U.S. soil Pb concentrations from Smith
et al. (2013a) are provided in Table 11-1. Regional studies of soil Pb, including Richardson et al. (2014).
which provides information on temporal trends of Pb concentrations in northeastern forest floor soils, and
Mackowiak et al. (2021). which surveyed soil and vegetation Pb concentrations in four counties in
southern California, are summarized in Section 11.2.3.
11-12
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4CP
3CP
2Cf
Source: Smith et al. (2013a).
Figure 11-1 Locations of the 4,857 soil sampling sites included in the U.S.
Geological Survey North American Soil Geochemical Landscapes
Project conducted from 2007 to 2010.
The 2006 Pb AQCD and 2013 Pb ISA reported representative Pb concentrations in fresh surface
water (median 0.50 (ig Pb/L. range 0.04 to 30 (ig Pb/L) and freshwater sediments (median 28 mg Pb/kg
dry weight, range 0.5 to 12,000 mg Pb/kg dry weight) in lotic systems in the United States based on a
synthesis of National Water Quality Assessment (NAWQA) data (U.S. EPA. 2013, 2006). Another
analysis of the NAWQA data set provides additional detail to the prior 2006 Pb AQCD analysis by
stratifying the summary of Pb concentrations in freshwater sediment by land use within river basins
(Horowitz and Stephens, 2008). The baseline freshwater sediment concentration, comprising
measurements taken in low-population areas only, is reported to have a median of 20 mg Pb/kg with a
range of 2 to 200 mg Pb/kg. Land-use categories for agricultural, cropland, pasture, forested and
rangeland sites are reported in Table 11-1. A more recent survey of Pb concentrations in freshwater
sediment found higher concentrations in Atlantic rivers (mean 110 mg Pb/kg) compared with Pacific and
Gulf of Mexico rivers (means of 19 and 32 mg Pb/kg, respectively) (Horowitz et al., 2012). This observed
Base map from U.S. Geological Survey data
Lambert Conformal Conic projection
600 KILOMETERS
600 MILES
11-13
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spatial variation in freshwater sediment Pb concentrations is likely driven by higher historical population
density and industrial activity associated with Pb emissions in the eastern United States compared with
the central and western regions of the country. Mahler et al. (2006) dated sediment cores and reported a
decline in Pb concentrations in sediment deposited between the 1970s and the 1990s, which corresponds
to the phasing out of widespread use of leaded gasoline. One additional regional survey of dissolved and
particulate Pb in fresh surface water was identified for inclusion in this ISA. In a study of water quality in
eight Texas rivers, Jiann et al. (2013) identified elevated particulate and dissolved Pb near areas with
greater anthropogenic influence and noted that Pb concentrations were decreased downstream of dams
and reservoirs, where slow-moving water causes suspended Pb to settle into sediment. Summary statistics
of the rivers included in Jiann et al. (2013) are included in Table 11-1. Additional information on
temporal trends observed in aquatic ecosystems is summarized in Sections 11.3.3 and 11.4.3.
No new surveys in coastal areas of the United States measuring dissolved Pb in saltwater or Pb in
saltwater sediment were identified for inclusion in this Pb ISA, although concentrations measured from
1984 to 1987 are included in Table 11-1 to provide additional information on Pb concentrations in
saltwater sediment (Cantillo and O'Connor. 1992). The 2013 Pb ISA (U.S. EPA. 2013) reported saltwater
dissolved and sediment Pb concentrations from studies dated 1977 to 1990 summarized in Sadiq (1992).
which reports a global range of 0.6 to 1,050 mg Pb/kg in saltwater sediment, although the authors noted
that the maximum value reported was observed in an Australian inland saltwater lake. Observations from
only one U.S. saltwater sediment study were reported in Sadiq (1992). in which Pb concentrations
ranging from 13.4 to 52.8 mg Pb/kg from Puget Sound were recorded. Sadiq (1992) remains the only
study identified for inclusion in the ISA in which global dissolved saltwater Pb concentrations are
reported. Excluding observations from inland seas, open-ocean concentrations of dissolved Pb ranged
from 0.01 to 4.8 |ig Pb/L. Pb measurement methods have developed substantially in the last few decades,
and measurements of dissolved Pb from older studies may be less accurate than those measured using
modern methods. Table 11-1 summarizes the information available on concentrations of dissolved and
sediment Pb observed in U.S. saltwater aquatic ecosystems.
Information on Pb concentrations observed in regional surveys of U.S. biota at sites located far
from significant modern point sources of Pb have been collated in Table 11-1. The included surveys
provide a range of reference values which may provide context for Pb concentrations observed in similar
species and ecosystems. The Western Airborne Contaminants Assessment Project (WACAP) is the most
comprehensive database on contaminant transport and depositional effects in U.S. sensitive ecosystems
(U.S. EPA. 2013; Blett. 2010; Landers et al.. 2010). although it only covers locations in the western part
of the country. The project aimed to assess the locations where atmospheric pollutants were accumulating
due to deposition in remote ecosystems in the western United States and identify the most likely sources
of the identified pollutants. Pb (and other pollutants) was measured in sediment, snow, water, lichen, and
fish at eight western U.S. national parks. For species sampled across multiple national parks, Pb
concentrations in biota in terrestrial and aquatic ecosystems surveyed in this project were reported in the
2013 Pb ISA and are included in Table 11-1.
11-14
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Recent regional surveys of Pb in terrestrial ecosystems published in the peer-reviewed literature
include Anna's hummingbirds (Calypte anna) surveyed in the coastal, valley and Sierra Nevada foothills
regions of northern California (Mikoni et al.. 2017) and cormorants (Phalacrocorax brasilianus) sampled
from two colonies in Lake Livingston and Richland Creek, Texas (Mora et al.. 2021). A summary of
feather Pb concentrations observed in each of these studies is included in Table 11-1. The study of Anna's
hummingbirds is unique in its investigation of bioaccumulation of metals in a nectar-feeding bird species.
The sources of Pb measured in hummingbird organs and feathers were not determined in this study, but
the authors listed absorption from food sources including plant and insect species, particularly those
living in urban environments, as the most likely routes of exposure (Mikoni et al.. 2017). Mora et al.
(2021) investigated the interaction between location and sex on Pb concentrations in cormorant feathers in
the Trinity River watershed in Texas and found no statistically significant effect for either variable.
A study of seven species of earthworms at nine sampling sites in the northeastern United States
was conducted alongside a concurrent soil survey that characterized the properties of the soil from which
the earthworm specimens were collected (Richardson et al.. 2015b). This study provides an example of
how Pb from many sources in environmental media is distributed throughout a regional terrestrial
ecosystem, observed in both earthworms and the soil they inhabit. Earthworm Pb concentrations were
found to be poorly correlated with the Pb concentrations in the soil horizons they were sampled from,
which is explained in part by the selectiveness of earthworms' feeding and the unknown fraction of
bioavailable Pb in the measured soil Pb. Concentrations measured in earthworm species sampled in
Richardson et al. (2015b) are summarized in Table 11-1.
Surveys of mussels (Mytilus sp.) and oysters (Crassostrea spp.) have been used to monitor Pb
concentrations in coastal ecosystems. The U.S. national Mussel Watch project (discussed in aquatic
temporal trends Section 11.4.3) has served as a biomonitoring network for Pb in coastal U.S. ecosystems
(Kimbrough et al.. 2008). An analysis of 2003-2006 Mussel Watch data including oysters (Crassostrea
gigas, Crassostrea virginica) and mussels (Mytilus edulis) identified a higher range of Pb concentrations
on the east coast of the United States relative to the west coast of Canada (Shiel et al.. 2012) (Table 11-1).
In this study, isotopic analysis and the covariance of cadmium (Cd) and zinc (Zn) were used to identify
the sources of Pb. Higher concentrations of Pb in the oysters and mussels on the east coast are attributed
to coal combustion and industries such as smelting and steelmaking.
The Large River Monitoring Network of the Biomonitoring of Environmental Status and Trends
(BEST-LRMN) surveyed fish from nine U.S. river basins from 1995 to 2004. This survey is the most
recent national-scale survey of Pb concentrations observed in biota in freshwater aquatic ecosystems, with
results summarized in two studies. Hinck et al. (2008) measured species-dependent Pb concentrations in
whole-fish common carp (Cyprinus carpio) and black bass (Micropterus spp.), and Hinck et al. (2009)
presented average Pb concentrations measured across species including black bass, white bass (Morone
spp.), catfish (Ictaluridae), northern pike (Esox lucius), northern pikeminnow (Ptychocheilus
oregonensis), burbot (Lota lota), trout (Salmonidae), pikeperch (Sander spp.), and goldeneye (Hiodon
11-15
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cilosoides) (Table 11-1; summary statistics of Pb observations are presented with each included species
combined). The BEST-LRMN survey is the most comprehensive study of bioaccumulation of Pb in fish
from U.S. ecosystems.
11.1.4 Concepts Related to Ecosystem Effects of Pb
Organism exposure and response to Pb in the various environmental media must be considered in
the context of the ecosystem. An ecosystem is a functional unit consisting of living organisms, their
nonliving environment, and the interactions within and between them (Allwood et al.. 2014). The
boundaries of what could be called an ecosystem are somewhat arbitrary, depending on the focus of
interest or study. Thus, the extent of an ecosystem may range from very small spatial scales to, ultimately,
the entire biosphere (Allwood et al.. 2014). Ecosystems can be natural, cultivated, or urban (U.S. EPA.
1986) and may be defined on a functional or structural basis. "Function" refers to the suite of processes
and interactions among the ecosystem components that involve energy or matter. Examples include water
dynamics and the flux of trace gases such as rates of photosynthesis, decomposition, and nutrient cycling.
Biotic or abiotic structure may also define an ecosystem. Abiotic structure includes climatic and edaphic
components. Biotic structure includes species abundance, richness, distribution, evenness, and
composition measured at the population, species, community, ecosystem, or global scale. A species (for
eukaryotic organisms) is generally defined by a common morphology, genetic history, geographic range
of origin, and ability to interbreed and produce fertile offspring. A population consists of interbreeding
groups of individuals of the same species that occupy a defined geographic space. Interacting populations
of different species occupying a common spatial area form a community (Barnthouse et al.. 2008).
Community composition may also define an ecosystem type, such as a pine forest or a tall grass prairie.
Pollutants can affect the ecosystem structure at any of these levels of biological organization (Suter et al..
2005).
When an ecological receptor encounters Pb, this metal may affect uptake processes and/or
interact with biological membranes. In some instances, depending on the form of Pb and prevailing
environmental chemistry, Pb is taken up by biota which can then lead to a biological response. The
alteration of cellular ion status (including disruption of Ca2+ homeostasis, altered ion transport
mechanisms, and perturbed protein function through displacement of metal cofactors) appears to be the
major unifying mode of action underlying all subsequent modes of action in plants, animals, and humans
(U.S. EPA. 2013). Molecular mechanisms linked to oxidative stress may induce DNA damage and
generation of reactive oxygen species (ROS), leading to protein modification, lipid peroxidation, and
altered enzyme response. Initial perturbations such as cytological or biochemical changes associated with
Pb exposure may cascade up to effects at higher levels of biological organization (i.e., from the
subcellular and cellular level through the individual organism and up to ecosystem-level processes). In
this ISA, biochemical (e.g., enzymes, stress markers) endpoints at the suborganism level of biological
organization are grouped under the broad endpoint of "physiological stress." Organism-level effects
11-16
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include reproduction, growth, and survival. These endpoints also have the potential to alter population,
community, and ecosystem levels of biological organization (Suter et al.. 2004). Causality determinations
for ecological effects of Pb in the 2013 Pb ISA used biological scale as an organizing principle to
summarize effects on vegetation, invertebrates and vertebrates in terrestrial, freshwater and saltwater
environments. The same approach is applied in this appendix, focusing especially on the organism-level
endpoints of reproduction, growth, survival, and effects on ecosystems.
In natural environments, where many variables that may impact the effects of interest are left
uncontrolled, partitioning the variability of responses and attributing observed effects to Pb unequivocally
is difficult. The presence of confounding factors that is characteristic of field observational studies is also
compounded by high natural variability in organismal genetics and in abiotic seasonal, climatic, water
chemistry or soil-related factors (U.S. EPA. 2015). In natural environments, modifying factors affect Pb
bioavailability and toxicity, and considerable uncertainties are associated with generalizing effects
observed in controlled studies to effects at higher levels of biological organization. Differences in
environmental chemistry may enhance or inhibit uptake of metal from the environment, thus creating a
spatial patchwork of environments that are at greater risk than other environments. Similarly, organisms
vary in their degree of adaptation to, or tolerance of, the presence of metals. Generally, the correct
attribution of effects to Pb is expected to be most challenging in studies that examine its effects on entire
ecosystems, as they incorporate all of the ecological interactions among the various populations and all of
the chemical and biological processes that affect Pb bioavailability (Section 11.1.6). The fundamental
principles of how metals interact with organisms and ecosystems are described in detail in U.S. EPA's
Framework for Metals Risk Assessment (U.S. EPA. 2007).
11.1.5 Ecosystem Services
In general, both ecosystem structure and function play essential roles in providing goods and
services. "Ecosystem services" refers to the concept that ecosystems provide benefits to humans, directly
or indirectly (Costanza et al.. 2017). and that ecosystems produce socially valuable goods and services
deserving of protection, restoration, and enhancement (Bovd and Banzhaf. 2007). The concept of
ecosystem services recognizes that human well-being and survival are not independent of the rest of
nature, but rather that humans are an integral and interdependent part of the biosphere (Costanza et al..
2017). In some cases, ecosystem services analysis can result in attaching monetary values to ecosystem
outcomes. However, because ecosystem services are often public goods, their benefits can be difficult to
monetize. Although the ecosystem services literature has expanded since the 2013 Pb ISA, there are few
publications that specifically link an ecological effect attributed to Pb to a change in an ecosystem
service. No new studies were identified that explicitly address Pb effects on ecosystem services associated
with terrestrial, freshwater, or saltwater systems.
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11.1.6 Bioavailability
As discussed in prior AQCDs and sections 6.6.3 (terrestrial), 6.4.4 (freshwater) and 6.4.14
(saltwater) of the 2013 Pb ISA (U.S. EPA, 2013), bioavailability is a key concept for understanding Pb
effects on the biotic components of ecosystems. U.S. EPA defines bioavailability as "the extent to which
bioaccessible metals absorb onto, or into, and across biological membranes of organisms, expressed as a
fraction of the total amount of metal the organism is proximately exposed to (at the sorption surface)
during a given time and under defined conditions" (U.S. EPA, 2007). This section presents a general
overview of bioavailability and introduces modifying factors and models to estimate bioavailability.
Chemical and biological modifying factors affecting bioavailability and subsequent toxicity to biota are
considered in more detail in the following sections: Section 11.2.2 (terrestrial), Section 11.3.2
(freshwater) and Section 11.4.2 (saltwater).
Bioavailability increases with the amount of Pb available as free Pb ions (U.S. EPA, 2013).
Factors affecting bioavailability and subsequent effects of Pb on biota include chemical factors that can
be quantitatively linked to toxicity. In soils, these include but are not limited to pH, cation exchange
capacity (CEC) and organic carbon (OC) content. In aquatic systems, water chemistry conditions
including hardness, pH, alkalinity and colloidal or dissolved OC (DOC) as well as the presence of other
metals affect the availability of Pb at sites of action on biological membranes. In saltwater, higher levels
of ions additionally affect Pb bioavailability. In sediments, Pb bioavailability may be influenced by the
presence of other metals, sulfides, iron (Fe) and manganese (Mn) oxides, and physical disturbance. In
addition to chemical factors, biological factors (see Section 7.2.3, (U.S. EPA, 2006) and Section 6.4.9,
(U.S. EPA, 2013)) affect bioavailability; however, they are more difficult to link quantitatively to
toxicity.
The bioavailability of a metal is also dependent upon the fraction of metal that is bioaccessible.
As stated in the Framework for Metals Risk Assessment (U.S. EPA, 2007), the bioaccessible fraction of a
metal is the portion (fraction or percentage) of environmentally available metal that interacts at the
organism's contact surface and is potentially available for absorption or adsorption by the organism. The
framework states that "the bioaccessibility, bioavailability, and bioaccumulation properties of inorganic
metals in soil, sediments, and aquatic systems are interrelated and abiotic (e.g., OC) and biotic
(e.g., uptake and metabolism) modifying factors determine the amount of an inorganic metal that interacts
at biological surfaces (e.g., at the gill, gut, or root tip epithelium) and that binds to and is absorbed across
these membranes. A major challenge is to consistently and accurately measure quantitative differences in
bioavailability between multiple forms of inorganic metals in the environment." A conceptual diagram
presented in the Framework for Metals Risk Assessment (U.S. EPA, 2007) summarizes metals
bioavailability and bioaccumulation in aquatic, sediment, and soil media (Figure 11-2).
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Bioaccessible Fraction (BF)a:
Percent soluble metal ion
concentration relative to total
metal concentration (measured in
solution near biomembrane)
Relative Bioavailability (RBA)b:
Percent adsorbed or absorbed
compared to reference material
(measure of membrane dynamics)
vailability
Absolute Bioavailability (ABA)c
Percent of metal mass absorbed
internally compared to external
exposure (measures systemic
uptake/accumulation)
Bioaccumulation of metal
==========-» Effects
¦~I Soluble species A
*1 Snl.ihlc tncrict R [—^ ^ .^Uptate~^
1 1 V ^vgfficiericii-^
Soluble species C 1 *
Benign
accumulatic/n
Internal
Transport
and
Distribution
Toxicoloical
accumulation
Site of
Toxic
Action
aBF is most often measured using in vitro methods (e.g., artificial stomach), but should be validated by in vivo methods.
bRBA is most often estimated as the relative absorption factor, compared with a reference metal salt (usually calculated on the basis
of dose and often used for human risk, but can be based on concentrations).
°ABA is more difficult to measure and used less in human risk; it is often used in ecological risk when estimating bioaccumulation or
trophic transfer.
Source: ERG (2004) and U.S. EPA (2007).
Figure 11-2 Conceptual diagram for evaluating bioavailability processes and
bioaccessibility for metals in soil, sediment, or aquatic systems.
The development and continued refinement of models that predict toxicity by incorporating
factors affecting bioavailability in aquatic systems have advanced the field of risk assessment for metals
(Adams et al.. 2020). The physicochemical composition of the receiving water determines the
bioavailability and thus the toxicity of metals to aquatic organisms. Therefore, aquatic bioavailability
models must incorporate the effects of influential aspects of water chemistry on metal toxicity. The biotic
ligand model (BLM) is a mechanistically based model for predicting the toxicity of single metals under a
large range of water chemistry conditions that considers complexations with inorganic ligands and
competition of active free metal ions with other cations, such as calcium (Ca) and magnesium (Mg), for
the site of action (i.e., biotic ligand) (Nivogi and Wood. 2004; Paquin et al.. 2002; Di Toro et al.. 2001). It
predicts both the bioaccessible and bioavailable fraction of Pb in the aquatic environment and can be used
to estimate the importance of environmental variables such as DOC in limiting uptake by aquatic
organisms (Alonso-Castro et al.. 2009). The U.S. EPA-recommended freshwater ambient water quality
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criteria (AWQC) for copper (Cu) are based on the BLM. Deforest et al. (2017) proposed a BLM-based
freshwater aquatic life criteria for Pb (Section 11.3.5).
Another recent approach to describing and predicting bioavailability and subsequent toxicity of
metals in aquatic environments are empirically based multiple linear regression (MLR) models, which
take into consideration a wide range of endpoints and water chemistry parameters from large empirical
toxicity data sets (Brix et al.. 2020). Since the 2013 Pb ISA, some studies have focused on further
evaluating the suitability of bioavailability models for predicting the chronic toxicity of Pb to aquatic
biota (Deforest et al.. 2017; Nvs et al.. 2016b; Nvs et al.. 2014). while others have explored the
development and evaluation of bioavailability models to predict the acute and chronic toxicity of metals
mixtures, in which Pb is a component (Nvs et al.. 2017; Farley et al.. 2015; Santore and Ryan. 2015). A
detailed consideration of the advancements in metal bioavailability modeling approaches is beyond the
scope of this ISA. A recent U.S. EPA report titled Metals Cooperative Research and Development
Agreement (CRADA) Phase I Report: Development of an Overarching Bioavailability Modeling
Approach to Support U.S. EPA 's Aquatic Life Water Quality Criteria for Metals evaluates and compared
BLM and MLR approaches for the purpose of updating the AWQC for Pb and other metals and advocated
for the use of MLR models over the BLM in future AWQC for metals (U.S. EPA. 2022). A review of the
current status and regulatory applications of metal bioavailability models is provided in (Mebane et al..
2020). For historical perspective, refer to (Adams et al.. 2020) and see (Brix et al.. 2020) for empirical
bioavailability model development.
In terrestrial environments, predicting responses to Pb exposure under field conditions from
exposure-response experiments that use soluble salts of Pb to spike study soils has met longstanding
difficulties, chiefly because of the differences in the many interacting determinants of bioavailability and
the difficulty of identifying and quantifying those interactions. Ports et al. (2021) recently suggested that
two bioavailability corrections to the results of those experiments may be sufficient: one to adjust for
percolation and aging, and the other to correct differences in toxicity that arise from differing soil
properties. The authors demonstrated the derivation of predicted no-effect concentrations (PNEC)
according to the European Registration, Evaluation, Authorisation and Restriction of Chemicals
(REACH) Regulation European Parliament and Council (2006) using the two corrections and data that
conformed to the REACH requirements.
11.1.7 Risk Screening Tools
Risk assessors have developed tools for identifying the concentrations of Pb in environmental
media that are at or below the thresholds for effects on ecological receptors. The following sections
present ecological screening criteria available for evaluating Pb in atmospheric deposition, soil, water,
sediment, and biota.
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11.1.7.1 Critical Loads for Atmospheric Deposition
The critical load concept is widely used as an organizing principle to relate atmospheric
deposition to ecological endpoints that indicate impairment (Pardo ct al.. 2011: Bobbink et al., 2010;
Porter and Johnson, 2007). The definition of a critical load is "a quantitative estimate of an exposure to
one or more pollutants below which significant harmful effects on specified sensitive elements of the
environment do not occur according to present knowledge" (Nilsson and Grennfelt, 1988). No recently
published critical loads for Pb from terrestrial ecosystems in the United States were identified for this
ISA. Several critical load studies from Europe reviewed in the 2013 Pb ISA (de Vries and Groenenberg,
2009; Hall et al., 2006; Morselli et al., 2006) and a recent review study (Koptsik and Koptsik. 2022) noted
uncertainties inherent in a critical load approach to Pb risk assessment, such as soil type, critical
concentration of dissolved metal, adsorption coefficients of exposed soils, combined effects of different
metals in multimetal mixtures and the influences of a changing climate. Since the 2013 Pb ISA, critical
load studies for atmospheric deposition for aquatic systems have largely focused on eutrophication and
acidification associated with nitrogen (N) deposition, with no detailed assessments for Pb in freshwater or
coastal areas in Europe (RoTAP, 2012) or the United States. In the literature search for the current
assessment, no published critical loads for atmospheric deposition of Pb were identified for U.S. inland or
coastal waters.
11.1.7.2 Soil Screening Levels
Developed by U.S. EPA, ecological soil screening levels (Eco-SSLs) are maximum contaminant
concentrations in soils that are predicted to result in little or no quantifiable effect on terrestrial receptors.
The Pb Eco-SSL was completed in March 2005 and has not been updated since. Values for terrestrial
birds, mammals, plants, and soil invertebrates are 11, 56, 120 and 1,700 mg Pb/kg soil (dry weight),
respectively. These conservative values were developed so that contaminants that potentially present an
unacceptable hazard to terrestrial ecological receptors are reviewed during the risk evaluation process
while removing from consideration those that are highly unlikely to cause substantive effects. The studies
considered for the Eco-SSLs for Pb and detailed consideration of the criteria for developing the Eco-SSLs
are provided in the 2006 Pb AQCD (U.S. EPA, 2006). Preference is given to studies using the most
bioavailable form of Pb to derive values. Soil concentrations protective with respect to avian and
mammalian exposure through diet are calculated by first converting dietary concentration to dose (mg/kg
body weight per day) for a critical study, then using food (and soil) ingestion rates and conservatively
derived uptake factors to calculate a soil concentration that would result in unacceptable dietary doses.
This approach frequently results in Eco-SSL values below the average background soil concentration
(U.S. EPA, 2005a, 2003), as is the case with Pb for the birds Eco-SSL. Sample et al. (2019) used a re-
analysis of some of the early studies included in the 2005 derivation of the avian Eco-SSL to propose a
new value.
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11.1.7.3 Ambient Water and Sediment Quality Criteria
AWQC represent surface water concentrations intended to be protective of aquatic communities,
including recreationally and commercially important species. The most recent AWQC for Pb were
developed in 1984 by the U.S. EPA Office of Water, which employed empirical regressions between
observed toxicity and water hardness to develop hardness-dependent equations for acute and chronic
criteria for the protection of aquatic biota (U.S. EPA. 1985a). These criteria are published pursuant to
Section 304(a) of the Clean Water Act and provide guidance to states and tribes to use in adopting water
quality standards for the protection of aquatic life and human health in surface water. The AWQC for Pb
for aquatic life are expressed as a criterion maximum concentration (CMC) for acute toxicity and criterion
continuous concentration (CCC) for chronic toxicity (U.S. EPA. 2009. 1985a). In freshwater, the CMC is
65 |ig Pb/L and the CCC is 2.5 |ig Pb/L at a hardness of 100 mg/L.
The current U.S. EPA AWQC for Pb in freshwater, published in 1984, are hardness-based and the
chronic criteria were developed based on the acute-to-chronic ratio due to the lack of chronic toxicity tests
in freshwater biota at that time. Since the AWQC for Pb were first published, additional acute and chronic
toxicity data has become available and better characterization of factors that influence Pb bioavailability
including development of a BLM for Pb. In view of this information, several researchers have proposed
updated approaches for WQC derivation for this metal. Taking into account the range of surface water
chemistry across the United States and the inclusion of newer toxicity data, Deforest et al. (2017)
proposed a BLM-based acute Pb criteria range from 18.9 to 998 |ig Pb/L and chronic BLM-based Pb
criteria range from 0.37 to 41 |ig Pb/L for freshwater (Section 11.3.5). The lowest criteria were for water
with low DOC (1.2 mg/L), pH (6.7) and hardness (4.3 mg/L as calcium carbonate [CaCOs]), and the
highest criteria were for water with high DOC (9.8 mg/L), pH (8.2) and hardness (288 mg/L as CaCOs).
Compared to the current U.S. EPA AWQC for freshwater, the number of genera with acute toxicity data
increased from 10 to 32, and the number with chronic toxicity increased from 4 to 13, which enabled the
proposed chronic criteria to be based on bioassay data rather than an acute-to-chronic ratio. Furthermore,
DOC and pH are represented in BLM; these water quality factors have a significant influence on Pb
bioavailability and toxicity along with hardness and other water characteristics (Adams et al.. 2020).
In comparison to the freshwater chronic criteria proposed by Deforest et al. (2017). Pb effect
thresholds to protect 95% of freshwater species calculated by Van Sprang et al. (2016) for seven selected
European freshwater scenarios were between 6.3 |ig Pb/L and 31.1 |ig Pb/L, based on chronic toxicity
datasets. There were several differences in development of the European thresholds for chronic Pb
toxicity compared with U.S. EPA guidelines, including the use of the 10% effect concentration (ECio)
rather than EC20 chronic toxicity data, selection of species mean values rather than genus mean values and
consideration of toxicity data for plants and algae in combination with bioavailability models to derive
effect thresholds. Furthermore, the range of water chemistries considered did not include the high
bioavailability conditions evaluated in (Deforest et al.. 2017).
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For freshwater sediment, U.S. EPA guidance has not changed since the 2006 Pb AQCD, and a
summary of the guidance is provided here. U.S. EPA has recommended sediment quality benchmarks for
Pb that, although not truly regarded as criteria, are concluded to be protective of benthic organisms.
Although sediment quality criteria have not been formally adopted, U.S. EPA has published an
equilibrium partitioning procedure for developing sediment criteria for metals (U.S. EPA. 2005b). For
freshwater sediment, the two approaches first summarized in the 2006 Pb AQCD, based on either bulk
sediment or equilibrium partitioning, continue to be used and refined. The first approach is based on
empirical correlations between metal concentrations in bulk sediment and associated biological effects to
derive threshold effect concentrations (TEC) and probable effects concentrations (PEC) (MacDonald et
al.. 2000). The TEC/PEC approach incorporates numeric guidelines to compare bulk sediment
concentrations of Pb. The equilibrium partitioning approach published by U.S. EPA for developing
sediment criteria for metals (U.S. EPA. 2005b) considers bioavailability by relating sediment toxicity to
the porewater concentration of metals. The amount of simultaneously extracted metal (SEM) is compared
with the metals extracted via acid volatile sulfides (AVS), since metals that bind to AVS (such as Pb)
should not be toxic in sediments where AVS occurs in greater quantities than SEM. The SEM approach
was further refined in the development of the sediment BLM (Di Toro et al.. 2005). An equilibrium
partitioning sediment benchmark for cationic metals, including Pb, was derived by Burgess et al. (2013).
The mechanistic-based sediment quality guideline was developed from the equilibrium partitioning
theory, in which the dissolved phase of Pb in sediment interstitial water serves as a surrogate for
bioavailable Pb. In the equation to derive the equilibrium partitioning sediment benchmark (Equation 11-
1), AVS are subtracted from SEMs to determine the amount of metal that could become bioavailable. The
equation takes into account interactions with both AVS and OC.
SEM - AVS
— A"ocFCV
Equation 11-1
The final chronic value (FCV) (|ig/L) in the equation is calculated with the following formula
(Equation 11-2) using a conversion factor (CF) for Pb in freshwater (Equation 11-3). The FCV for Pb in
saltwater is 8.1 |ig/L.
p£y _ £pre1.273(Jn(harcfness))-4.705T
Equation 11-2
CF = 1.46203 — 0.145712 {ln(hardness))] Equation 11-3
The most recent aquatic life AWQC for Pb in saltwater were released in 1984 (U.S. EPA. 1985a)
by U.S. EPA's Office of Water. These criteria are published pursuant to Section 304(a) of the Clean
Water Act and provide guidance to states and tribes to use in adopting water quality standards for the
protection of aquatic life and human health in surface water. The AWQC for Pb are currently expressed as
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CMC for acute toxicity and CCC for chronic toxicity (U.S. EPA. 2009). In saltwater, the CMC is
210 |ig Pb/L and the CCC is 8.1 |ig Pb/L.
Since the most recent update of the U.S. EPA AWQC for saltwater, there are considerably more
acute and chronic toxicity data available for saltwater organisms, which reduce uncertainties related to Pb
toxicity and regulatory thresholds. For example, the 1985 CCC for saltwater was calculated based on
acute-to-chronic ratios from freshwater biota (Church et al.. 2017; U.S. EPA. 1985a). The U.S. EPA's
guidelines for derivation of AWQC indicate that when there are sufficient data, comparison of toxicity
data sets from different taxa using species sensitivity distributions (SSDs) can be performed to estimate
criteria values through a probabilistic approach and to set the level of protection (U.S. EPA, 1985). The
minimum diversity required to develop SSDs has historically precluded this method for saltwater biota
due to lack of toxicity data. Using ECio acute toxicity data from sensitive early lifestages of 13 species
representing 7 taxa (phytoplankton, polychaetes, bivalves, crustaceans, echinoderms, chordates, fish)
inhabiting Atlantic European coastal ecosystems, Duran and Beiras (2013) derived an acute saltwater
quality criterion for Pb of 25.3 |ig Pb/L from SSD. This value, derived from the lower end of the 95%
confidence intervals of the 5th percentile of the SSD, is intended to protect 95% of species in 95% of
cases. Church et al. (2017) proposed an updated saltwater acute criterion of 100 |ig Pb/L and chronic
criterion of 10 |ig Pb/L based on genus mean toxicity values following U.S. EPA methodology (U.S.
EPA. 1985b) (Section 11.4.5). Church et al. (2017) derive regulatory values using species sensitivity
distributions (SSDs) for which some of the toxicity values are from data sources not included in the ISAs,
which are also not necessarily published in peer-reviewed literature (i.e., unpublished reports, university
theses, memoranda).
Methods for establishing marine sediment guidelines and sediment quality values used globally
were recently reviewed by Birch (2018). Sediment quality values for U.S. waters were generally in the
range of the sediment quality threshold values reported by MacDonald et al. (1996). with a threshold
effects level of 30 mg Pb/kg and a probable effects threshold of 112 mg Pb/kg. A low effects threshold of
46.7 mg Pb/kg sediment and median effects threshold of 218 mg Pb/kg sediment were the sediment
quality guidelines developed for the National Oceanic and Atmospheric Administration (NOAA) National
Status and Trends Program (NOAA. 1999).
11.2 Terrestrial Ecosystems
11.2.1 Summary of New Information on Effects of Pb in Terrestrial
Ecosystems and Causality Determination Update Since the 2013 Pb ISA
Since the 2013 Pb ISA (U.S. EPA. 2013). evidence has continued to accrue for many of the
effects of Pb on terrestrial ecosystems reported in the ISA and previous U.S. EPA assessments. This
additional support includes investigations of effects on species and communities that had not been
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studied, but none of the additional evidence is sufficient to change any of the conclusions for terrestrial
ecosystems that were reached at the time. There are no changes to existing causality determinations
for terrestrial biota or ecosystems from the 2013 Pb ISA (Table 11-2).
Additional observational studies published after the 2013 Pb ISA (U.S. EPA, 2013), many of
which were anthropogenic environmental gradient studies, have linked Pb exposure and effects on
microbial community structure (e.g., abundance, diversity) and function (e.g., enzyme activities,
respiration rates). Many found mixed (negative, positive, or null) relationships between total or
bioavailable Pb soil concentration and the abundance of bacterial and fungal taxa. It remains difficult to
disentangle the effects of Pb exposure on microbial communities from the effects of other soil
contaminants using anthropogenic environmental gradients, as other heavy metals and soil
physicochemical properties are significantly correlated with soil Pb concentration, and many of these
factors also influence microbial processes.
Studies published since the 2013 Pb ISA (U.S. EPA, 2013) continue to support previous findings
that plants tend to sequester larger amounts of Pb in roots as compared with shoots and that there are
species, ecotype, and cultivar-dependent differences in the uptake of Pb from soil and the atmosphere, and
in translocation of sequestered Pb. In the 2013 Pb ISA (U.S. EPA, 2013), the body of evidence was
sufficient to conclude there is a causal relationship between Pb exposure and plant physiological stress
and a causal relationship between Pb exposure and plant growth. Evidence was inadequate to determine
causal relationships between Pb exposure and both plant survival and plant reproduction. Recent studies
have continued to demonstrate various deleterious physiological effects of Pb exposure on plants,
particularly oxidative stress. Strong uncertainties remain regarding the concentrations at which these
effects would be observed in the environment. Recent studies have examined the protective effects of
mycorrhizae and of some plant nutrients when added in excess of the minimal requirements of the plants.
In terrestrial invertebrates, the evidence reviewed in the 2013 Pb ISA (U.S. EPA, 2013) was
sufficient to conclude that there is a causal relationship between Pb exposure and decreased survival and
between Pb exposure and reproductive and developmental effects, a likely causal relationship between Pb
exposure and decreased growth, neurobehavior effects and physiological stress, and the evidence is
inadequate to conclude that there is a causal relationship between Pb exposure and hematological effects.
Evidence collected since then provides additional support for the effects of Pb exposure on organismal
and suborganismal responses including a decrease in survival, and decreased growth and fecundity.
Recently published studies on physiological responses to Pb include decreases in protein and lipid content
and increases in malondialdehyde (MDA) in earthworms. Acetylcholinesterase (AChE) activity decreased
in response to Pb in snails and honeybees while the effects on protein, glycogen, other enzymes, and
glutathione-S-transferase (GST) responses were variable depending on the site or species examined.
Several new studies quantified behavioral changes to Pb exposure in bees. Evidence also suggests that in
earthworms, Pb exposure can have lasting effects on growth even postexposure on earthworms and slow
the time to maturation. Pb exposure delayed onset of the breeding season and shortened duration in
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isopods, as well as influenced mate selection in fruit flies. Evidence published after the 2013 Pb ISA
(U.S. EPA, 2013) includes new organisms as well as modifying factors of organism response such as
habitat, exposure history, and seasonality.
Effects of Pb commonly observed in terrestrial vertebrates include decreased survival, and
reproduction, as well as effects on development and behavior (U.S. EPA, 2006). The 2013 Pb ISA (U.S.
EPA, 2013) also provided evidence for Pb effects on hormones and other biochemical variables. In the
2013 Pb ISA (U.S. EPA, 2013) the body of evidence was sufficient to conclude that there is a causal
relationship between Pb exposure and reproductive and developmental effects, and between Pb exposure
and hematological effects, and a likely causal relationship between Pb exposure and decreased survival,
physiological stress, and neurobehavioral effects for terrestrial vertebrates. The evidence was inadequate
to conclude that the relationship between Pb exposure and growth is causal for terrestrial vertebrates.
Studies published since the 2013 Pb ISA provide additional evidence for effects on suborganism- and
organism-level endpoints, and specifically on hematological and physiological endpoints, but they do not
affect determinations of causality. New studies have expanded upon the relationship between Pb exposure
and ALAD activity by adding more species of birds, amphibians, and mammals to the evidence base.
More evidence of oxidative stress has been gathered, as well as evidence of effects on corticosterone
levels and immunity in birds. Literature since the 2013 Pb ISA continues to add to evidence relating to
reproductive effects at both the organism and suborganism levels including effects on lifetime breeding
success and some specific secondary sexual traits. New studies of behavioral effects included increased
aggression in mockingbirds.
Systematic studies of the validity of using results of Pb salt-addition experiments for estimating
effects of Pb exposure under field conditions have continued since the 2013 Pb ISA. As previously,
experiments showed that the form of Pb, pH, CEC, OC, Fe and Mn oxides, percolation, aging, and soil
composition are all strong modifiers of toxicity. Recent studies demonstrated additional interactions
among those variables and showed that their effects are at times mediated by additional variables such as
salinity. Those studies continue to support the conclusion that data from exposure-response experiments
in terrestrial environments conducted using spiking of soils with soluble salts of Pb, are unlikely to
generate accurate estimates of effects in contaminated natural environments. However, Ports et al. (2021)
suggested that two corrections to the results of exposure-response experiments conducted with additions
of soluble salts of Pb to soil may be sufficient to derive predicted no-effect concentrations (PNEC)
according to the European REACH Regulation European Parliament and Council (2006).
In the 2013 Pb ISA (U.S. EPA, 2013) the body of evidence was sufficient to conclude that there
is a likely causal relationship between Pb exposure and terrestrial-community and ecosystem effects.
Some new evidence of the effects of Pb at higher levels of biological organization is available, but it is
insufficient to change the determination of causality. Species interactions between tree species and their
pests, and between herbaceous plants and nectar robbers, worms and lepidopteran consumers were among
the new community and ecosystem endpoints for which effects of Pb were observed. Several studies
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found negative relationships between Pb concentration along a pollution gradient and aspects of
invertebrate community structure, specifically in soil mites, potworms, insect communities on kale and
nematodes. Although evidence for effects on growth, reproduction, and survival at the individual
organism level and in simple trophic interactions makes the existence of effects at higher levels of
organization likely, direct evidence is relatively sparse and difficult to quantify. The presence of multiple
stressors, especially including other metals, continues to introduce uncertainties in attributing causality to
Pb at higher levels of organization.
Table 11-2 Summary of Pb causality determinations for terrestrial plants,
invertebrates, and vertebrates
Level Effect Terrestrial3
2013 Pb ISAb
2024 Pb ISA
Community and Ecosystem
Community and Ecosystem Effects
Likely Causal
Likely Causal
Reproductive and Developmental
Effects - Plants
Inadequate
Inadequate
Reproductive and Developmental
Effects - Invertebrates
Causal
Causal
Population-
level
Endpoints
Reproductive and Developmental
Effects - Vertebrates
Causal
Causal
Growth - Plants
Causal
Causal
Organism-level
Growth - Invertebrates
Likely Causal
Likely Causal
Responses
Growth - Vertebrates
Inadequate
Inadequate
Survival - Plants
Inadequate
Inadequate
Survival - Invertebrates
Causal
Causal
Survival - Vertebrates
Likely Causal
Likely Causal
Neurobehavioral Effects -
Invertebrates
Likely Causal
Likely Causal
Neurobehavioral Effects - Vertebrates
Likely Causal
Likely Causal
Hematological Effects - Invertebrates
Inadequate
Inadequate
Hematological Effects - Vertebrates
Causal
Causal
Suborganismal
Responses
Physiological Stress - Plants
Causal
Causal
Physiological Stress - Invertebrates
Likely Causal
Likely Causal
Physiological Stress - Vertebrates
Likely Causal
Likely Causal
Conclusions were based on the weight of evidence framework for causal determination in Table II of the ISA Preamble (U.S. EPA.
2015) 2013 Pb ISA (U.S. EPA. 2013).
bEcological effects observed at or near Pb concentrations measured in soil, sediment, and water in Table 6-2 of the 2013 Pb ISA
were emphasized and studies generally within one to two orders of magnitude above the reported range of these values were
considered in the body of evidence for terrestrial systems (Section 6.3.12) (U.S. EPA. 2013).
Previous AQCDs and the 2013 Pb ISA identified uncertainties with regard to the contribution of
Pb from current deposition to soil Pb concentration and subsequent toxicity to terrestrial biota, as opposed
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to historic contributions. Historic Pb from gasoline and other sources as well as Pb from current air and
non-air sources is present in terrestrial systems and moves through the different environmental media
(e.g., soil, sediment, water, biota) confounding source apportionment. The contribution of atmospheric Pb
to specific sites is not clear (U.S. EPA, 2013). Furthermore, as stated in the 2013 Pb ISA, many factors,
including species and various soil physiochemical properties, interact strongly with Pb concentration to
modify effects. In terrestrial ecosystems, where soil is generally the main component of the exposure
route, Pb aging is a particularly important factor, and one that may be difficult to reproduce
experimentally. Without quantification of those interactions, characterizations of exposure-response
relationships would likely not be transferable outside of experimental settings (U.S. EPA, 2013). Key
uncertainties with regard to Pb effects in terrestrial ecosystems in the last review included the
uncertainties expected from widening the scope of inference from controlled laboratory studies to
conditions in natural environments, where many modifying factors affect Pb bioavailability and toxicity.
This also applies when going from studies at low levels of biological organization to effects at higher
levels. Conversely, it is difficult to partition the variability of responses and to attribute observed effects
to Pb unequivocally in natural environments, where many variables that may impact the effects of interest
are left uncontrolled. The presence of confounding factors that is characteristic of field observational
studies is also compounded by high natural variability in organismal genetics and in abiotic seasonal,
climatic, water chemistry or soil-related factors (U.S. EPA, 2015). For instance, available studies on
community and ecosystem-level effects are usually from contaminated areas where Pb concentrations are
much higher than typically encountered in the environment and where multiple contaminants are present.
Studies that characterize bioavailability, uptake, bioaccumulation, and effects of Pb in terrestrial
ecosystems or that decrease uncertainties identified in the prior Pb NAAQS review and were published
since the 2013 Pb ISA (literature cutoff for inclusion in the 2013 Pb ISA was September 2011) are
presented throughout the following sections. Brief summaries of conclusions from the 1977 Pb AQCD
(U.S. EPA. 1977). 1986 Pb AQCD (U.S. EPA. 1986). 2006 Pb AQCD (U.S. EPA. 2006) and 2013 Pb
ISA (U.S. EPA. 2013) are included where appropriate. Recent research on the bioavailability and uptake
of Pb into terrestrial biota including plants, invertebrates and vertebrates is presented in Section 11.2.2.
Environmental concentrations in terrestrial biota and ecosystems in the United States at different locations
and over time are discussed in Section 11.2.3. The toxicity of Pb to terrestrial biota (Section 11.2.4) is
followed by data from exposure-response studies (Section 11.2.5). Responses at the community and
ecosystem levels of biological organization are reviewed in Section 11.2.6.
11.2.2 Factors Affecting Bioavailability, Uptake and Bioaccumulation and
Toxicity in Terrestrial Biota
Long-range atmospheric transport of Pb and natural rock weathering are the primary sources of
Pb in natural systems away from anthropogenic point sources. Non-urban terrestrial ecosystems
potentially affected by Pb deposition include natural forests, managed forests, grasslands, pastures, and
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cropland. Once deposited, Pb can be resuspended into the air or transferred among other environmental
media. Pb atmospheric inputs into terrestrial ecosystems include direct deposition as well as resuspension
and transport of historically deposited Pb from nearby roads and contaminated soils (Appendix 1
https://assessments.epa.gov/isa/document/&deid=359536). In terrestrial systems, Pb is distributed
between biota, soil, and soil porewater. Mobility of Pb into biotic components of the ecosystem is a
function of the chemical speciation of Pb and subsequent bioavailability. Bioavailability of Pb in soils
(Section 11.1.6) depends on local soil physicochemical properties including pH, CEC, organic matter
(OM), inorganic compounds, salinity, clay content and aging. Uptake experiments with terrestrial plants
and invertebrates generally show increases in Pb uptake with increasing Pb concentration in the medium
but with strong effects from several interacting factors (U.S. EPA, 2013, 2006). Below, factors that
affect bioavailability of Pb in terrestrial systems are summarized along with information that advances
understanding of Pb uptake in terrestrial biota since the 2013 Pb ISA.
11.2.2.1 Factors Affecting Bioavailability of Pb in Terrestrial Biota
The 2013 Pb ISA described the bioavailable fraction of Pb in soil as being strongly dependent on
the fraction of Pb dissolved in soil porewater, which is primarily controlled by processes related to
partitioning of Pb between liquid and solid phases: (1) solubility equilibria; (2) adsorption-desorption
relationship of total Pb with inorganic compounds (e.g., oxides of aluminum (Al), Fe, silicon (Si), Mn;
clay minerals); (3) adsorption-desorption relationship reactions of dissolved Pb phases on soil OM; (4)
pH; (5) CEC; and (6) aging (U.S. EPA, 2013). The 2013 Pb ISA summarized studies that confirmed the
role each of these six factors plays in the sequestration and release of Pb in soil porewater (U.S. EPA,
2013). Total metal loading is described by the 2013 Pb ISA as the most influential factor controlling
adsorption and desorption, with higher concentrations of Pb corresponding to an overall decrease in the
fraction of Pb adsorbed to organic and inorganic surfaces (U.S. EPA, 2013). However, even as the
adsorbed fraction decreases with increasing metal loading, the rate of that decrease and the fraction of
adsorbed Pb will vary considerably between different soil types. This variability can be attributed to
differences in soil physicochemical properties, pH, CEC, OM, inorganic compounds, salinity, and aging.
These physicochemical properties are based on soil forming factors: climate, organisms, parent material,
relief, time, and anthropogenic input. Soils that differ in these factors will subsequently have different
physicochemical properties and considerable differences in the environmentally available fraction of Pb.
In addition, although predictions of bioavailability and toxicity based on environmentally available
fractions using extractable or porewater concentrations are still generally supported, evidence from recent
studies suggests that there may be limitations in predicting toxicity from environmentally available
concentrations represented as either porewater or calcium chloride (CaCl2)-extractable concentrations
(Lanno et al„ 2019; Bur et al„ 2012; Paugct et al„ 2011).
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11.2.2.1.1 pH and Cation Exchange Capacity
The 2013 Pb ISA cited a study conducted by Smolders et al. (2009) wherein models of metal
bioavailability calibrated from 500+ soil toxicity tests on plants, invertebrates and microbial communities
indicated pH and CEC were the most important factors governing both metal solubility and toxicity.
Recent literature confirms these findings and continues to highlight the important influence that pH and
CEC have on Pb bioavailability. To identify the main physicochemical factors controlling Pb
bioavailability in earthworms, Tang et al. (2018) conducted toxicity experiments on earthworms exposed
to 13 soils with low-level Pb contamination and varying physicochemical properties. Bioaccumulation
factors (BAFs) were calculated for each of the 13 soils and stepwise MLR and path analyses were used to
assess the relationships between soil physicochemical properties and BAFs. Results showed that the Pb
BAFs of earthworms in soils with pH <5.5 were higher than those in other soils. OC, pH, and total Pb in
soil were identified as the most important physicochemical parameters controlling Pb bioavailability. The
authors concluded that their results confirmed that low pH increases Pb mobility, which promotes uptake
and subsequent bioaccumulation (Tang et al.. 2018). Romero-Freire et al. (2015) demonstrated the
important influence of pH on bioavailability by measuring Pb toxicity to plants and bacteria exposed to
aqueous extracts from seven soils with different physicochemical properties. Both Pb solubility and
toxicity were significantly correlated with pH, CO3 and OC. Of the seven soils that were assessed, sandy
acidic soil with the lowest pH was associated with the highest extractable Pb concentration and the lowest
half maximal effect concentration (EC50) value for the plant bioassay. Wiiavawardena et al. (2015)
investigated the relationship between soil properties and relative bioavailability in swine exposed to 11
different soils spiked with Pb. Freundlich partition coefficients (Kd) were calculated for each soil, and
stepwise regression analysis was used to evaluate the relationships between different soil properties and
relative bioavailability as well as Kd partition coefficients. Regression models showed that pH and clay
content were the most influential soil properties, accounting for 85% and 54% of variability in Kd and the
relative bioavailability of Pb, respectively. Lanno et al. (2019) examined the effects of physicochemical
properties on the toxicity of Pb to two different soil invertebrates, collembolans (Folsomia Candida) and
earthworms (Eisenia fetida), in seven different soils spiked with Pb salts at varying concentrations. EC50
values varied considerably amongst the different soil types, ranging from 35 to 5,080 mg/kg for
earthworms and 389 to >7,190 mg/kg for collembolans. BAFs were also calculated for earthworms and
varied with a >10-fold range across the different soil types. Effective CEC (eCEC) and soil properties
related to eCEC including total C, exchangeable Ca and Mg and clay content had a significant effect on
both Pb toxicity and bioaccumulation as well as the toxicity thresholds EC10 and EC50 in earthworms.
However, there were no correlations between soil properties and Collembola toxicity threshold
concentrations. The authors suggested that reduced toxicity in Collembola may be attributed to species-
dependent differences in Pb uptake across epidermal surfaces, specifically the sclerotized cuticles of
collembolans may reduce the uptake of Pb2+ across epidermal surfaces, limiting uptake to intestinal
absorption from ingestion of soil porewater. The study also assessed whether variability in toxicity values
was better explained using exposure estimates based on environmental available fractions (measured as
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Pb2+ in porewater or as total dissolved Pb in porewater) rather than total Pb in soil. The results showed
greater variation in EC50 values based on environmentally available fractions compared with EC50 values
based on total Pb soil concentrations. These results combined with significant correlations between
earthworm endpoints and eCEC, but not pH, may suggest that eCEC reduces Pb uptake by cation
exchange of Pb2+ in both clay and OC coupled with competition for uptake between multiple cations at
the surface of the earthworm epithelium. The competition for cation uptake at the epithelial surface may
also extend to H+, which may help explain why toxicity thresholds were not correlated with pH.
Additional explanations for greater toxicity variability in porewater may also be due to unexplained
chemical interactions between Pb2+ and soil porewater as well as the physiological mechanisms of
earthworm absorption and metabolism. Similar results were reported in a study that examined Pb and Cd
bioavailability in soils located in the vicinity of a smelter; uptake rate constants of Pb in earthworms were
significantly greater at higher pH. Giskaet al. (2014) suggested that higher pH may be associated with a
decrease in competition between heavy-metal ions and H+ ions for binding sites on biotic ligands.
11.2.2.1.2 Organic Matter and Inorganic Compounds
The 2013 Pb ISA described the significant roles that both organic and inorganic soil constituents
play in immobilizing Pb and decreasing bioavailability. Surfaces of both OM and inorganic materials
(clays and sesquioxide minerals) contain negatively charged functional groups, which serve as sites of Pb
adsorption. In addition, Pb can form immobile precipitates with CO3, phosphate and sulfate that may also
be present in soil porewater. Shaheen and Tsadilas (2009) noted that soils with higher clay content, OM,
total CaCO , equivalent and total free sesquioxides also exhibited higher total Pb concentration, indicating
that less Pb had been taken up by resident plant species.
While recent studies confirm findings from the 2013 Pb ISA regarding the roles of OM and
inorganic surfaces in Pb immobilization, they also suggest that OM is capable of increasing or decreasing
Pb mobility. Shahid et al. (2012) reviewed the role of humic substances on Pb phytoavailability and
toxicity and concluded that the overall role of humic substances in Pb bioavailability is complex due to
the heterogenous nature of humic substances and varying soil physicochemical properties. Depending on
both of these factors, humic substances may exist as dissolved OM (DOM) capable of binding free Pb2+ in
soil porewater, as solid constituents with high adsorption affinity for Pb or as DOM capable of increasing
the extractable and bioavailable fractions of Pb. de Santiago-Martin et al. (2014) used bioassays with
romaine and iceberg lettuce grown in calcareous Mediterranean soils with low levels of OM that were
spiked with Pb, Cu, Cd and Zn to assess the contribution of soil physicochemical properties toward
bioavailability. CO3, OM and fine mineral fractions accounted for 85% of the variance in bioavailability,
and OM was the most important variable explaining Pb and Cd bioavailability patterns. However, OM
seemed to exert contrary effects on Pb and Cu bioavailability. At lower concentrations of the metals, OM
and bioavailability were negatively correlated, but a positive correlation was observed at higher
concentrations. The authors suggested that differences in the role OM had at different concentrations may
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be attributed to competitive binding between Pb and Cu onto humic acids, resulting in a larger
bioavailable fraction at higher concentrations due to saturation of binding sites on humic acids (de
Santiago-Martin et al.. 2014). Similar results of the contradictory role that OM may have on
bioavailability were reported by Zeng et al. (2011). whereby OM was observed to have a positive
correlation with ethylenediaminetetraacetic acid (EDTA)-extractable chromium (Cr), Cu, Fe, Mn, Pb and
Zn, and both positive and negative correlations with concentrations in rice straw grown in the
contaminated soils. Pauget et al. (2011) evaluated the influence of pH, OM and clay content on chemical
availability and bioavailability of Pb to land snails (Cantarens cispersus) exposed to nine contaminated
soils, each differing by a single characteristic (pH, OM, or clay content). The results demonstrated that an
increase in both pH and OM decreased Pb bioavailability to snails. However, clay did not have a
significant influence. It is worth noting that the clay mineral used for this assessment was kaolinite.
Kaolinite is 1:1 clay with no interlayer spaces and only external exchange sites at the edges of tetrahedral
and octahedral sheets. As a result, kaolinite has a low CEC compared to other clay minerals. Other clay
minerals (2:1) with both external and internal exchange sites in interlayer spaces may have had more
influence on bioavailability. The authors of the study acknowledged this limitation and other studies have
conveyed the important role that clay can have in decreasing metal mobility (de Santiago-Martin et al..
2013). Pauget et al. (2016) investigated the contributions of soil and lettuce to bioavailability in garden
snails (Cantareus cispersus) and the influence of soil properties, pH, and OM on the contribution of each
source. Results indicated that soil contributed to 90% of Pb bioavailability in snails exposed to both soil
and lettuce, and increasing OM content further increased the contribution by an additional 6%. The
authors suggested that increasing OM may have also resulted in increased DOM, which may have
increased the soluble fraction of Pb through formation of DOM-Pb complexes in soil solution. An
additional explanation suggested for the increased bioavailability in soil with higher OM may be an
increase in ingestion rate caused by a decrease in nutrients following the addition of OM.
11.2.2.1.3 Salinity and Aging
In addition to the physicochemical properties described above, Pb mobility and bioavailability
can also be influenced by salinity. Application of CaCk MgCl or NaCl salts to field-collected soils
containing 31 to 2,764 mg Pb/kg increased the proportion of the mobile metal fraction. As the strength of
the salt application was increased from 0.006 to 0.3 M, the proportion of released Pb increased from less
than 0.5% to over 2% for CaCb and from less than 0.5% to over 1% for MgCl (Acosta et al.. 2011).
However, the majority of salinity-induced effects occurred in soils containing less than 500 mg Pb/kg,
and the proportion of released Pb decreased with increasing total soil Pb concentrations. Recent literature
continues to show that laboratory soils spiked with Pb2+ salts, which are commonly used in toxicity
studies, may overestimate toxicity in corresponding field-contaminated soils (Figure 11-3) due to lack of
aging as well as increases in salinity and acidification that occur after the soil has been spiked with Pb2+
salts (Smolders et al.. 2015). Smolders et al. (2015) compared Pb toxicity between three groups of soils:
(1) aged 5 years, leached and pH-corrected, (2) leached and pH-corrected, and (3) freshly spiked soils
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with no leaching or pH corrections. Leaching, pH correction and aging after spiking reduced toxicity to
plant, microbial and invertebrate receptors by a factor of 8 (median value) based on ECio values. ECio
values were often near background levels for freshly spiked soils, but after leaching, pH correction and
5 years of aging, the majority of ECio values were above 1,000 mg/kg. The authors concluded that salinity
stress, rather than acidification or aging, is the main factor explaining increased Pb toxicity in freshly
spiked and unleached soils and suggested that researchers performing future toxicity tests consider
spiking soils with lead monoxide (PbO) fine powder rather than PbCh salt to exclude confounding salt
effects. PbO fine powder would also be more representative of Pb that contaminates soil through
atmospheric deposition. Similar results demonstrating the importance of aging were reported by Zalaghi
and Safari-Sincgani (2014). In the study, soils were spiked with 0, 600, 1,200 and 1,800 mg/kb Pb as lead
nitrate (Pb(NC>3)2), and the environmentally available fraction of Pb and microbial toxicity were measured
at select time increments across a 90-day period. The concentrations of Pb in the environmentally
available fraction and microbial toxicity showed a considerable decrease over the 90-day period of the
study. The authors concluded that this decrease in bioavailability was due to the transfer of Pb into CO3
and residual fractions that occurred as a result of aging. Similar results demonstrating a decrease in Pb
bioavailability following soil aging were reported by (Zhang and Van Gestel. 2019a).
20 30 40 60 100 200 300 500 1000 2000 4000 7000
Added Pb (mg Pb/kg soil)
Source: Smolders et al. (2015).
Figure 11-3 Change in toxicity expressed as relative responses (i.e., response
relative to the mean of the corresponding control soil) for three
different laboratory soil treatments: freshly spiked; spiked,
leached and pH-corrected; and spiked, leached and pH-corrected
with 5 years of aging.
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11.2.2.1.4 Biological Factors
The severity of Pb effects on terrestrial biota depends in part upon species differences in
metabolism, sequestration, and elimination rates. Because of the effects of soil aging and other
bioavailability factors discussed above, in combination with differing species assemblages and biological
accessibility, ecosystems may also differ in their sensitivity and vulnerability to Pb. The 2006 Pb AQCD
and 2013 Pb ISA reviewed these factors, including nutritional factors, soil aging and bioavailability.
Sensitivity to Pb exposure was found to vary widely among terrestrial species, even among closely related
organisms. It was noted that in many species of birds and mammals, dietary factors can exert significant
influence on the uptake and toxicity of Pb. Since the 2013 Pb ISA, new information on soil aging has
further expanded understanding of factors that modify soil bioavailability under natural conditions.
To disentangle the effects of salinity, acidification, and aging on the sensitivity of microbial
communities, plants, and invertebrates to Pb, Smolders et al. (2015) conducted an experiment in which
toxicity to these groups was tested in soils spiked with Pb2+ salts, leached and aged. Uncontaminated
soils were collected from grasslands and agricultural lands in Spain, the United Kingdom and Belgium
and were exposed to 0, 250, 500, 1,000, 2,000, 4,000 or 8,000 mg Pb/kg using PbCh Some of the soil
was set aside (treatment: freshly spiked), while the rest was incubated for a week, leached using artificial
rainwater and pH-corrected to maintain soil pH within 0.2 pH units within each Pb concentration using
CaO (treatment: leached and pH-corrected). Five years prior to spiking soils with PbCh, additional soils
were exposed to the same Pb gradient using Pb(NC>3)2 and stored in perforated pots which were left
outdoors to age. After 5 years, pH was corrected using CaO (treatment: aged, leached and pH-corrected).
Soil solution Pb concentration, i.e., porewater Pb concentration, increased in a dose-dependent manner
with spiked soils, followed by leached soils and finally aged soils containing the least soil solution Pb
(except in aged soils from Spain). Toxicity was then tested in microbial communities, earthworms (E.
fetida), Collembola (/¦'. Candida), tomato (Lycopersicon esculentum) and barley (Hordeum vulgare).
Toxicity was highest in freshly spiked soils (mean ± S.E., ECsofor all organisms tested: 2,300 ± 145 mg/
kb Pb), followed by leached and pH-corrected soils (6,500 ± 750 mg Pb/kg) and then aged soils (>10,000
mg Pb/kg); however, the effects of leaching with pH correction and aging with pH correction were
inconsistent among organisms and toxicity tests. Depending on the origin of the soil, leaching and pH
correction reduced toxicity based on ECio values by a factor of 1.9-2.3 compared with freshly spiked
soils, while aging and pH correction reduced toxicity by a factor of 2.7-13. Microbial activity (potential
nitrification rate, substrate-induced nitrification, and respiration rate), invertebrate reproduction and plant
growth were negatively correlated with total soil Pb concentration, porewater Pb concentration, Pb2+ ion
activity and porewater ionic strength. With the exception of E. fetida reproduction, these factors were
positively correlated with soil pH. Given porewater ionic strength had the strongest influence on toxicity
across all tested organisms, the authors suggest that salt stress may modify the toxicity of Pb, as
acidification and aging were unable to explain variation in toxicity.
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11.2.2.1.5 Summary
In summary, studies published since the 2013 Pb ISA continue to substantiate the important role
that soil geochemistry plays in sequestration or release of Pb and its bioavailability to organisms.
Environmentally available concentrations, measured either in soil porewater or as extractable Pb, are
generally still a useful predictor of bioavailability, although predictions cannot be transferred between
experiments with soluble salts of Pb and field conditions. pH is still considered the most important factor
influencing the concentration of Pb in this fraction due to its important role in Pb solubility. However,
several studies have reported results that suggest limitations in using the environmentally available
fraction to predict bioavailability and toxicity. These studies suggest species-dependent uptake and
metabolism mechanisms as well as other soil physicochemical properties that may be involved in
chemical interactions between soil porewater and biological receptors should be considered. Inorganic
compounds, including clay minerals and sesquioxides, particularly Fe and Mn oxides are still considered
to play important roles in Pb sequestration, and CEC is still a reliable measure of a soil's ability to sorb
and exchange cations, which is an important function for Pb sequestration. The role of OM in Pb
sequestration and mobility remains complex. Depending on the nature of the OM and soil
physicochemical properties, Pb may bind to solid OM surfaces, decreasing Pb mobility. Alternatively,
OM may enhance Pb release into soil solution through the formation of Pb-DOM complexes or following
OM decomposition. Studies published since the 2013 Pb ISA also continue to highlight limitations in
using laboratory soils spiked with Pb salts to predict toxicity in field-contaminated soils. Many of these
studies have demonstrated that the use of Pb2+ salts in laboratory soils without adequate leaching, pH
correction and aging greatly affects Pb bioavailability and leads to overestimating the toxicity that would
be expected to occur in field-contaminated soils with similar concentrations of Pb.
11.2.2.2 Uptake and Bioaccumulation in Terrestrial Plants
Studies published since the 2013 Pb ISA continue to support previous findings that plants tend to
sequester larger amounts of Pb in roots as compared with shoots, and that there are species-, ecotype-, and
cultivar-dependent differences in uptake of Pb from soil and the atmosphere and translocation of the
sequestered Pb (U.S. EPA, 2013, 2006, 1977). Further, many species of plants accumulate heavy metals
in environments with extreme soil concentrations and are therefore used for phytoremediation at such
sites. Although occasional phytoremediation studies may be informative with respect to the mechanisms
of Pb uptake and tolerance, most do not add further evidence with respect to the effects of atmospheric
Pb. The same applies regarding mosses and lichens as biomonitors of atmospheric Pb. Despite Pb not
being a plant nutrient, it is taken up from soils through the symplastic route, the same active ion transport
mechanism used by plants to take up water and nutrients and move them across root cell membranes
(U.S. EPA, 2006). As with all nutrients, only the proportion of a metal present in soil porewater is
directly available for uptake by plants. In addition, soil-to-plant transfer factors in soils enriched with Pb
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have been found to better correlate with bioavailable Pb soil concentration, defined as diethylenetriamine
pentaacetate-extractable Pb, than with total Pb concentration (U.S. EPA, 2006).
Previous reviews (U.S. EPA, 2013, 2006, 1977) noted that terrestrial plants accumulate
atmospheric Pb primarily via two routes: direct stomatal uptake into foliage and incorporation of
atmospherically deposited Pb from soil into root tissue, followed by variable translocation to other
tissues. It was recognized that most Pb taken up from soil remains in the roots and that distribution to
other portions of the plant is variable among species. Most of the Pb absorbed from soil remains bound in
plant root tissues either because (1) Pb may be deposited within root cell wall material or (2) Pb may be
sequestered within root cell organelles, which may be a protective mechanism for the plant. Studies since
the 2013 Pb ISA have generally confirmed that Pb taken up from soil largely remains in the roots (Naikoo
et al., 2019; Zhou et al., 2019; Zhou et al., 2015; Meiman et al., 2012; Rossato et al., 2012).
Previous findings have shown that Pb translocation to stem and leaf tissues does occur at
significant rates in some species, including some crops and herbaceous species (e.g., rattlebush,
buckwheat, Chinese cabbage, pak-choi, and water spinach). There is broad variability in uptake and
translocation among plant species, and interspecies variability has been shown to interact with other
factors such as soil type. These results indicate significant interspecies differences in Pb uptake, as well as
potential soil-dependent differences in Pb bioavailability (U.S. EPA, 2013).
Although exposures are often high, field studies carried out in the vicinity of Pb smelters and
other industrial point sources have determined the relative importance of direct foliar uptake and root
uptake of atmospheric Pb deposited in soils, with greater overall uptake corresponding to closer proximity
to the source (Angelova et al., 2010; Hu and Ding, 2009; Cui et al„ 2007). Hu and Ding (2009) concluded
that metal accumulation in some leafy greens grown in the vicinity of a smelter were greater in shoot than
in root tissue, which suggested both high atmospheric Pb concentration and direct stomatal uptake into
the shoot tissue. Similarly, evidence since the last review shows substantial accumulation of Pb in needles
in areas with high contributions of atmospheric Pb (Kandziora-Ciupa et al„ 2016; Gandois and Probst,
2012). Studies also noted a significant difference between Pb concentrations in washed and unwashed
leaves, indicating that aerial deposition and surface dust is likely a significant source of plant-associated
Pb (Ugolini et al., 2013; El-Rjoob et al„ 2008). Foliar Pb may include both incorporated Pb (i.e., from
atmospheric gases or particles) and surficial particulate Pb deposition. The plant may eventually absorb
the surficial component; however, the main importance of surficial Pb is its likely contribution to the
exposure of plant consumers or to leaf litter. The consideration of these Pb exposures to humans via
consumption of food crops is briefly discussed in Section 2.1.3 of Appendix 2.
Because of their long life spans, certain trees can provide essential information regarding the
sources of bioavailable Pb. A Scots pine (Finns sylvestris) forest in northern Sweden was found to
incorporate atmospherically derived Pb pollution directly from ambient air, accumulating this Pb in the
bark, needles, and shoots (Klaminder et al„ 2005). More recent studies have also shown that accumulation
in the bark of some species is a useful bioindicator of exposure to atmospheric Pb (Janta and Chantara,
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2017; Palowski et al., 2016). Metal content can also vary in relation to altitude as a result of long-range
transport. Korzeniowska et al. (2021) found that metal content in the moss (Pleurozium schreberi (Willd.)
Mitten) and in Norway spruce (Picea abies (L.) H. Karst) in the Tatra National Park in the Carpathian
Mountains of Poland was greater with increasing altitude.
Dendrochronology (tree-ring analysis) is an important tool for measuring the exposure of trees to
environmental Pb (Watmough. 1999). While effectiveness may vary by species investigated, tree-ring
studies reviewed in the previous AQCDs and ISAs showed that trees could be used as indicators of
increasing environmental Pb concentrations with time (U.S. EPA, 2013, 2006, 1977). Trees accumulate
and sequester atmospheric Pb in close correlation with the rate of point-source emissions (Guyette et al.,
1991). Studies published since the 2013 Pb ISA continue to demonstrate dendrochronology is a useful
tool for monitoring historical uptake of Pb into trees exposed to atmospheric or soil Pb (Sensula et al.,
2017; Dinis et al., 2016; Beramendi-Orosco et al„ 2013; Doucet et al., 2012) (Section 11.2.3).
In the 2013 Pb ISA (U.S. EPA, 2013), plant-associated arbuscular mycorrhizal fungi (AMF) were
found to protect the host plant from Pb uptake. Additional evidence indicates that the presence of AMF or
bacteria hosts can influence Pb accumulation in and alleviate Pb stress on plants. Inoculation of David's
mountain laurel (Sophorct dcividii, previously Sophorct viciifolia) with the AMF Funneliformis mossecte
resulted in lower concentrations of Pb in belowground and aboveground biomass (Xu et al., 2016a). S.
davidii seeds collected from around the Qiandongshan Pb and Zn mine in northwest China were grown in
pots receiving 0, 50, 500, or 1,000 mg Pb/kg (aqueous Pb(NC>3)2). Half of the pots with S. davidii plants
were inoculated with F. mossecte. After 4 months, mycorrhizal colonization, Pb accumulation, plant
height, diameter, aboveground and belowground biomass, and root characteristics were recorded
(Section 11.2.4.2). Vesicular, arbuscular, hyphal, and total root colonization of S. davidii decreased with
increasing Pb treatment. Both mycorrhizal and nonmycorrhizal plants showed increasing Pb content in
their roots and aboveground tissue in a dose-dependent manner, but belowground and aboveground Pb
concentrations were lower for mycorrhizal plants. Pb concentration in aboveground tissue of mycorrhizal
plants was 54%-66 % less Pb than that in nonmycorrhizal plants, while roots contained 15%—85 % less,
depending on Pb exposure. The root-to-shoot Pb concentration of mycorrhizal plants increased with Pb
exposure while nonmycorrhizal plant root-to-shoot concentration decreased with increasing Pb exposure,
suggesting that Pb was sequestered in the root following inoculation with F. mosseae. Furthermore,
transmission electron micrographs and X-ray microanalysis of S. davidii roots under different Pb and
mycorrhizal treatments suggested Pb in the cytoplasm was sequestered in the cell walls and vacuoles of F.
mosseae, while Pb was transported into the root cells and intracellular space of nonmycorrhizal plants.
Pot marigolds (Calendula officinalis) inoculated with Glomus mossea and G. intradices
accumulated more Pb relative to nonmycorrhizal plants, yet experienced greater fitness-(Tabrizi et al.,
2015). Calendula officinalis were grown in pots and received 0, 150, or 300 mg Pb/kg (aqueous
Pb(NC>3)2). Half of the plants were inoculated with a mixture of G. mossea and G. intradices. Root
colonization, Pb accumulation, plant growth, reproduction flavonoid contents and nutrients were
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analyzed. Root colonization decreased with increasing Pb exposure in a dose-dependent manner, as root
colonization in the control (0 mg Pb/kg) was 56% higher than in the high Pb treatment (300 mg Pb/kg).
Pb concentration in the roots and the shoots (mg Pb/plant) increased with increasing Pb exposure.
Inoculated Calendula officinalis had 10.3% more Pb in the roots compared with noninoculated plants,
while shoots of inoculated and noninoculated plants contained the same amount of Pb. The interaction
between Pb exposure and inoculation did not influence Pb uptake in aboveground or belowground
biomass.
In another example, the AMF Gigaspora margarita increased bioaccumulation of Pb but reduced
Pb-induced stress of silver banner grass (Miscanthus sacchariflorus) (Sarkar et al.. 2018). Miscanthus
sacchariflorus rhizomes and soil were collected from sites around the Ara River, Japan and placed in the
greenhouse. The collected soil contained 0.12 mg Pb/kg. Miscanthus sacchariflorus received 0, 100, or
1,000 mg Pb/kg additional Pb (aqueous), and half of the plants were inoculated with G. margarita. After
4 months, root colonization, bioaccumulation of Pb and plant growth, survival, hormones, enzymes,
nutrients, and chlorophyll content were characterized. Root colonization of M. sacchariflorus by G.
margarita decreased with increasing Pb concentration for both inoculated and noninoculated plants. The
Pb content of the belowground biomass of inoculated M. sacchari florus was higher than the Pb content of
noninoculated M. sacchariflorus. A similar pattern was observed for aboveground biomass, wherein
inoculated plants contained equal or higher concentrations of Pb than noninoculated plants.
Inoculation of black alder (Alnus glutonisa) by an actinobacteria, Frankia, affected Pb uptake in
roots and shoots (Belanger et al.. 2015). Alnus glutonisa seedlings were grown from seeds in the
laboratory and half were inoculated with Frankia alni (ACN14a), isolated from Alnus viridis ssp. crispa
in Quebec, Canada. Half of the inoculated and noninoculated control plants were exposed to Pb(NC>3)2
(0.10 mM). Pb exposure did not affect the nodule development of inoculated plants and Pb root
concentration was 4.3 times lower in roots and 6.3 times higher in shoots compared with inoculated A.
gultonisa not exposed to Pb.
In a recent study, (Gao et al.. 2021) reported that the type of mycorrhizal fungi (AMF versus
ectomycorrhizal fungi [EMF]) associated with seven tree species in an evergreen broadleaf forest in
China does not affect uptake of Pb from roots to leaves. Foliar and root tissues were collected and
analyzed for Pb concentrations as well as phosphorus (P), potassium (K), Ca, Mg, Fe, Mn, Cu, Zn,
strontium (Sr), total C, and total N. Elemental concentrations in the tree were analyzed according to their
mycorrhizal type (AMF versus EMF), plant organ (leaves versus roots) and an interaction term. Pb
concentrations were significantly higher in the roots compared with the leaves. The elemental Pb
concentrations between the roots and the leaves were uncorrelated for AMF-associated trees, EMF-
associated trees, and all species, suggesting that mycorrhizal type does not influence Pb uptake in the
roots or the leaves.
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11.2.2.3 Uptake and Bioaccumulation in Terrestrial Invertebrates
At the time of publication of the 2006 Pb AQCD (U.S. EPA, 2006), little information was
available regarding the uptake of atmospheric Pb pollution by terrestrial invertebrate species. Evidence in
the 2013 Pb ISA indicated that invertebrates, especially snails and earthworms, can accumulate Pb via
diet, exposure through soil, or from both exposure routes in the case of earthworms and snails. In the
2013 Pb ISA, snail Pb concentrations were reported to be lower than soil concentrations and uptake and
bioaccumulation were reported to be lower than the corresponding values for other metals (U.S. EPA,
2013). Exposure routes for soil organisms are through food consumption and soil exposure; soil variables,
such as pH and OM, influence uptake. Similarly, earthworm uptake is influenced by soil physicochemical
properties, genus, and the vertical position earthworms occupy within the soil profile (i.e., epigeic, epi-
endogeic, endogeic, anecic). Furthermore, earthworm activity in soil acts as a control on Pb
bioavailability and its uptake by earthworms, potentially other soil organisms, and plants. In addition to
providing supporting information on the uptake and availability of Pb to snails and earthworms, recent
literature has examined the bioavailability and accumulation of Pb with many other invertebrates
including lepidoptera, spiders and bees; in addition to soil factors (such as pH and OM), field
characteristics, organism sex and season may also influence uptake and accumulation. Since new
information has become available on organisms not discussed in previous assessments, these studies are
included despite being non-U.S. based.
11.2.2.3.1 Snails
In support of the 2013 Pb ISA conclusions regarding Pb uptake by snails, recent literature
continues to show snail tissue concentration is typically lower than soil concentration values. One recent
study found that when Pb was examined in soil, leaves, and snail tissues at increasing distance to metal
smelters, Pb in soils was, in general, highest closest to smelting plants and decreased with increasing
distance. Pb content in stinging nettle leaves (Urtica dioicci) followed the same general pattern of
decreasing Pb concentration with distance as did European land snail (Cepaea nemorctlis) digestive gland
tissue. The concentration in plant tissue was positively correlated with soil level, and snail tissue
concentration was positively correlated with plant tissue concentration. Patterns persisted over 4 months
of exposure. Nettles are the preferred food source of C. nemorctlis and exposure to Pb appears to be
primarily through consumption. While bioaccumulation factors (BAF) were not calculated, Pb
concentration in snail tissue was considerably lower than soil concentrations but was typically 2.5-3.5
times higher than plant tissues after 16 weeks of exposure (Boshoff et al., 2015) [see also (Nica et al.,
2012)1. However, one recent study suggests some snail species may be greater accumulators than others.
Vrankovict al. (2020) sampled Roman snails (Helixpomatia) foot muscles and hepatopancreas tissue
across a three-location urban gradient of soil Pb levels. Soil Pb varied from approximately 15 mg Pb/kg at
the reference (forest), approximately 30 mg Pb/kg at the medium pollution site and approximately
110 mg Pb/kg at the high pollution site. Foot muscle and hepatopancreas tissue concentration increased
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with increasing exposure levels. More Pb was stored in the hepatopancreas than the foot tissue, and
hepatopancreas levels were generally higher than soil contamination. BAF values were less than 1 for foot
muscle (0.47, 0.9, and 0.42) and greater than 1 for hepatopancreas tissue at the low and medium pollution
sites (1.61, 1.72, 0.76). The greater concentration found in the hepatopancreas indicates greater uptake via
food. Concentrations reported within snail tissues in this study were higher than those reported in studies
examining other snail species, suggesting uptake and accumulation are partly species-specific; see also
(Mleiki et al.. 2017).
New literature further supports that Pb uptake by snails is influenced by soil characteristics as
well as being dose- and duration-dependent. The concentration in the digestive gland of the green garden
snail (Cantareus apertus) increased with increasing exposure level after 1 week of exposure for low
(25 mg Pb/kg), medium (100 mg Pb/kg) and high (2,500 mg Pb/kg, dietary values measured) exposure
levels (Mleiki et al.. 2016). However, tissue concentration was not significantly greater in the
2,500 mg Pb/kg treatment compared with the 100 mg Pb/kg treatment. Similarly, after 8 weeks of
exposure, digestive gland tissue concentration was higher under Pb exposure compared with the control,
but the highest concentrations were found under the 100 mg Pb/kg exposure. An observational field study
examining the uptake and elimination kinetics of Pb by the common garden snail (Cantareus aspersus)
found soil Pb concentration (positive), CEC (positive) and soil OC content (negative) have a multivariate
effect on Pb bioavailability. Similarly, soil silt (positive), sand (positive) and OC content (negative)
modulate Pb uptake by snails (Pauget et al.. 2013b). In another study, soil Pb concentration was correlated
with Pb concentration in juvenile C. aspersus but when OC content and Al and Fe oxides were included
in the model, R2 increased from 0.37 to 0.56. The most polluted plots (i.e., plots with the highest Pb
concentration) did not have the highest Pb transfer to snails. OC content is known to influence metal
mobility and bioavailability for soil organisms (Pauget et al.. 2013a).
11.2.2.3.2 Earthworms
In the 2013 Pb ISA, studies of bioaccumulation of Pb in earthworms reported that many soil
physicochemical properties, including pH, OM and CEC, affect metal bioavailability for these organisms;
recent studies confirm these observations. Following 4 weeks in soil spiked with a solution of Pb (NOs^
(40, 250, 500, 1,000, 2,500 mg Pb/kg, nominal soil values, concentration in worms was measured),
juvenile E. fetida body Pb concentration increased with exposure concentration. BCFs ranged from 0.14
to 0.3, indicating either low bioavailability of Pb in the soil or low ability to accumulate Pb within tissues.
After 4 weeks of recovery (no Pb exposure), earthworm body Pb was significantly lower than the value at
the end of the exposure period but was still higher than the control and positively correlated with
exposure values (Zaltauskaite et al.. 2020). A study on native Eisenoides lonnbergi earthworms in
Maryland found E. lonnbergi can accumulate extraordinarily high levels of Pb, with a BAF of 83
recorded (Bever et al.. 2018). Accumulation was driven by soil Ca levels and indirectly by pH and clay
content, not by soil Pb content or availability. In acidic, low Ca soils, Pb uptake and accumulation is
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greater. Over soil Ca concentrations ranging from 49 to 1,695 mg Pb/kg, E. lonnbergi can maintain body
Ca concentrations between 4000 and 8,000 mg Pb/kg. Thus, even in Ca-poor soils, E. lonnbergi can
uptake enough needed Ca to maintain necessary body concentrations. The Ca BCF was 3.3 in high Ca
soils and 117 in low Ca soils. The Pb concentration factor was 1.02 in high Ca soils and 83 in low Ca
soils, suggesting Pb is absorbed by the Ca transport system, which is known to occur in vertebrates
(Bever et al.. 2018).
In E. fetida earthworms exposed to a range of soil Pb values from 125 to 350 mg Pb/kg across a
range of pH, Pb concentration in the worms was higher in low pH (<5.5) soils than in neutral or alkaline
soils with similar Pb concentrations (Tang et al.. 2018). Following 4 weeks in soil spiked with a solution
of Pb (NOs)2 (40, 250, 500, 1,000, 2,500 mg Pb/kg, nominal soil values, concentration in worms was
measured), juvenile E. fetida body Pb concentration increased with exposure concentration. BCF varied
from 0.14 to 0.3 indicating either low bioavailability of Pb in the soil or low ability to accumulate Pb
within tissues (Zaltauskaite et al.. 2020). After 4 weeks of recovery (no Pb exposure), earthworm body Pb
was significantly lower than at the end of the exposure period but was still higher than the control and
positively correlated with exposure values. In another study examining earthworm Pb concentrations,
BAFs in low pH soils were also higher than those in other soils but all BAFs were less than one
(Richardson et al.. 2015b). Soil Pb, OC, and pH together gave the best predictive model outcome on
earthworm Pb concentration. Earthworm ecotype can influence Pb tissue concentrations as well. Endogeic
and epigeic species were found to have higher Pb tissue concentration than epi-endogeic and anecic
earthworms (Richardson et al.. 2015b). A recent meta-analysis by Richardson et al. (2020) examined the
influence of soil concentration, soil characteristics, earthworm genus and ecotype on trace metal uptake.
They found soil concentration did not predict earthworm tissue concentration but ecophysiological group,
earthworm genus, metal source, exposure duration, and soil OM were important predictors.
In studies cited in the 2013 Pb ISA, earthworm feeding and burrowing behavior altered the
bioavailability, mobility, and uptake of Pb by earthworms and other soil biota. Recent studies further
elucidate the effects of earthworms on soil Pb processes. One study examined the decomposition of
Amvnthas agrestis and Liimbricus rubellus earthworms and the subsequent release of Pb in different
fractions within the soil column over 60 days (Richardson et al.. 2016b). Both species had similar Pb
tissue concentrations but due to the greater mass of A. agrestis added to experimental soils on a dry
weight basis, A. agrestis contributed a larger pool of Pb to the soil column. Leachate from both
earthworm treatments was significantly higher in Pb than leachate from control (no earthworm) soils.
Exchangeable Pb pools were greater under both earthworm treatments but only at days 7 and 21. By
day 60, there was only slightly more exchangeable Pb under the A. agrestis treatment compared with the
control. The stable Pb pool was greater under earthworm treatments across all sampling dates and the
majority of Pb under earthworm treatments was in the stable fraction. In a lab experiment using field-
collected polluted soils, Liimbricus terrestris earthworms were exposed to a high Pb-contaminated soil
(4,550 mg Pb/kg), a medium polluted soil (988 mg Pb/kg) and a low polluted soil (109 mg Pb/kg) for
28 days (Sizmur et al.. 2011a). By the end of the exposure period, earthworms had consumed less than
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2% of the bulk soil. Soil pH and water-extractable OC were higher in earthworm casts compared to
control soils. Earthworm casts had greater extractable and residual Pb pools and lower reducible pools.
Porewater from earthworm-inhabited highly contaminated soils had higher Pb concentrations compared
with control soils. Under the medium contamination Pb soils, there was more Pb2+ and inorganic Pb, but
less organic Pb compared with control soils. In low pollution earthworm soils, there was less Pb2+ but
more organic and inorganic Pb compared with control soils. While earthworms only processed a small
portion of the soil during the 28-day exposure treatment, the greater solubility of Pb from casts shows
earthworms can alter Pb bioavailability and is tied to the changes in pH and OC of the casts.
The effect of two invasive, but widespread, species of earthworms in northeastern U.S. forests
(Amvnthas agrestris and Lumbricus rubellus) on litter decomposition, metal exchange, and metal
bioaccumulation was examined in a laboratory experiment using forest floor material (collected from
New Hampshire) with and without earthworms (Richardson et al.. 2016a). Both species dwell at the soil
surface either in or just below the litter layer. Pb levels in forest floor and soil were approximately 26 and
16 mg Pb/kg, respectively. After 80 days, litter mass, percent carbon, and carbon mass were all lower in
the forest floor material when earthworms were present. Earthworm presence also resulted in lower
exchangeable Pb fraction concentrations but there was no difference between earthworm treatments and
control on the stable Pb fraction. Tissue concentration increased over time, with a BAF of 2.32 for A
agrestris after 80 days and 2.39 forZ. rubellus. The BAF for the exchangeable fraction was only 104.2
for A. ctgrestis and 88.3 forZ. rubellus. Both worms increased litter decomposition and carbon loss and
lowered the exchangeable Pb fraction. However, the stable Pb pool did not respond to earthworm
presence. Both earthworm species did accumulate Pb at greater concentrations than the forest floor and,
as mentioned by the authors, at levels higher than the maximum tolerable level approved for poultry and
mice feed, therefore posing a contamination risk to birds and small mammals. In an observational
experiment in New England forests, Pb soil concentrations and pools were examined in the presence or
absence of nonnative earthworms (Aporrectodea rosea, Dendrobaena octaedra, Aporrectodea
tuberculata were most common) (Richardson et al.. 2017). Like the previous study, Pb in New England
soils sampled in this study represent background Pb levels in an area of the country with a history of
metal enrichment via pollution. Pb concentration was lower in the Oa horizons at high abundance sites
compared with low abundance sites; however, within the A and E horizons, Pb was higher at high
abundance sites. Organic horizon Pb pools were negatively correlated with earthworm biomass, but total
soil Pb pools showed no relationship with earthworm biomass.
In a study that examined earthworm effects on the bioavailability and mobility of metals in soil,
leachates at the end of a 112-day exposure period had greater Pb concentration in the presence of Z.
terrestris earthworms (1.9 |ig Pb/L) compared with control soils (1.0 |ig Pb/L) (Sizmur et al.. 2011b). Pb
leachate from under Z. terrestris consisted of 98.4% Pb2+ as free ions and 0.9% as fulvic-acid-complexed
Pb compared with 95.7% and 4.0%, respectively, in control soil leachate. Soil pH was lower under all
earthworm species at the end of the experiment compared with the control. Perennial ryegrass (Lolium
perenne) was planted 28 days prior to soil sampling and harvested 21 days later. Ryegrass shoots had
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greater Pb concentrations when grown on columns with L. terrestris compared with grass grown in
control soils. The dry mass of plant shoots did not differ between treatments. The results showed
earthworms can increase Pb mobility and availability to plants, increasing sequestration. Over a 6-week
experiment, there was no effect of Pb on lettuce growth but when grown in soils with earthworms, lettuce
biomass increased with increasing concentrations (significantly higher at 3,730 mg Pb/kg concentration)
(Leveque et al.. 2014). Earthworms also increased lettuce Pb concentration but only at exposure
concentrations of 2,822 and 3,730 mg Pb/kg.
11.2.2.3.3 Other Invertebrates
For the 2013 Pb ISA, studies of bioavailability and uptake comprised earthworms, snails, and
arthropods including bees and beetles. Since the 2013 Pb ISA, new literature has examined additional
invertebrate groups including spiders, and butterflies. Pollen, honey, and bees from 16 honeybee {Apis
mellifera) apiaries were sampled twice a year for 2 years for Pb contamination across an urban-cultivated-
hedgerow-natural environmental gradient in France (Lambert et al.. 2012). Pb concentration in pollen was
influenced by sampling season but not by landscape characteristics. Thirty percent of honey samples were
below detection limits, and the rest had very low concentration values. Pb concentrations in honey from
apiaries surrounded by a hedgerow matrix were two times higher than those in other landscapes
measured, with honey from cultivated sites having the lowest concentrations (most were below the
detection limit). Pb in honey was higher in the 2009 season compared with the 2008 season. Pb
concentrations ranged from 0.001 to 1.896 mg Pb/kg in bees, from 0.004 to 0.798 mg Pb/kg in pollen and
from 0.004 to 0.378 mg Pb/kg in honey. Seasonality may influence bee Pb concentration, as levels were
higher in bees sampled during the June-October sampling period for one of the years studied. There was
no clear relationship of contamination between the three biological compartments (pollen, bees, honey).
In general, apiaries in urban and hedgerow locations had higher Pb contamination than apiaries in
cultivated or island landscapes. There was variation across the year, and contamination was typically
higher during the dry (summer) season. Honeybees are exposed to Pb contamination via direct contact
with Pb atmospheric deposition on flowers and through food contamination. Pb contamination patterns in
bees were similar to contamination levels in pollen, suggesting deposition contact contamination.
Seasonal differences may be explained by changes in floral availability.
Following 20 km-pollution gradients away from active Zn or metal smelters in Russia and
Poland, bumblebee {Bombus spp.) Pb levels (0.21-3.3 mg Pb/kg) and soil Pb levels (13.6—
814.2 mg Pb/kg) both decreased with increasing distances from the pollution source (Szentgyorgyi et al..
2011). In another study, bee body, bee bread, propolis, and honey Pb content was examined across
different geologic areas (Golubkina et al.. 2016). Sites included an unpolluted control located in the
Ribnitsa district in Moldavia (located away from industry or major highways), a selenium (Se)-deficient
area in the Henty province of Mongolia and the Voskresensk district of Moscow region, which is an area
of fertilizer production. Bee body Pb concentration was lowest at the unpolluted Moldavia location
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(0.51 mg Pb/kg), higher in Mongolia (0.94 mg Pb/kg), 0.97 mg/kg away from fertilizer production area
(Novoselki, Russia) and over 4 times higher near fertilizer production (2.16 mg Pb/kg, Lopatino). There
was a positive correlation between Pb content in bees and bee bread for the Lopatino and Moldavia sites.
Pb content in the propolis was highest in Mongolia (16.07 mg Pb/kg) and much lower in the other
locations (2.08, 1.52, 3.18 mg Pb/kg, Moldavia, Novoselki, Lopatino, respectively) and was not as closely
correlated with bee body content. Honey Pb content was low across all sites (approximately 0.2 mg Pb/kg
or less).
Wolf spiders (Lycosidae) are common ground-dwelling arachnids and are known to accumulate
metals. An observational study in Korea found that while Pb in soil did not differ by season
(31.13 mg Pb/kg averaged across seasons), Pb was significantly greater in spiders from an autumn brood
(7.83 mg Pb/kg) compared with that in a spring brood (1.52 mg Pb/kg). While overall BCF was below 1,
the difference in brood accumulation suggests that while spiders accumulate Pb at low levels, seasonality
may affect accumulation (Conti et al.. 2018). Jung and Lee (2012) measured Ariadna spider Pb
accumulation in Namibia in relation to uranium (U), Cu, and gold mines. Overall, Ariadna spiders do
accumulate heavy metals in relation to their environment (in this case burrowing spiders and sand
contamination), but Pb levels were higher in sand compared to the levels in spider bodies, indicating Pb is
not readily bioaccumulated.
In the common cutworm (Spodopterci liturci), Pb accumulation in body tissue generally increased
with increasing Pb exposure concentration across all development stages (Shu et al.. 2015). Larvae were
exposed to increasing Pb concentration via diet at 0, 12.5, 25, and 50 mg Pb/kg (dietary values) and larvae
were raised for five generations at each exposure concentration. Growth stage (larvae, pupae, adult), Pb
exposure concentration, and their interaction explained Pb accumulation, but generation did not (F1
versus F5), nor were there any significant interactions with generation. Within development stages, Pb
accumulation was highest during the 6th instar stage, second highest in adults, and lowest in pupae (Pb
accumulation was only significantly higher at 50 mg Pb/kg treatment for pupae and adults). Within 6th
instar larvae, Pb exposure and tissue type mattered but sex did not. Overall, Pb accumulated primarily
within the midgut, and overall gut accumulation (mid, fore, and hindgut) was greater than that in the
hemolymph, head, or body fat. Accumulation also increased with exposure. In a trophic uptake study, Pb
accumulation in the roots, stems and leaves of mulberry (Morns alba) increased with increasing soil Pb
exposure (0, 200, 400, 800 mg Pb/kg, dietary values) (Zhou et al.. 2015). In turn, Pb in silkworm
(Bombyx mori) larvae and moths as well as in feces and silk excretions increased with increasing Pb
content in the mulberry leaves (in response to increasing Pb in soil). However, larvae (0.63, 4.08, 5.74,
and 11.16 mg Pb/kg) and moths (0.6, 2.95, 4.39, 6.23 mg Pb/kg) had lower body content than leaves
(5.54, 41.79, 51.21, 60.26 mg Pb/kg) while Pb in feces was higher than that in leaves (9.85, 187.96,
230.44, 279.8 mg Pb/kg), indicating that while silkworms accumulate more Pb in response to increasing
exposure, Pb is not biomagnified, and the majority of Pb consumed is excreted instead.
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A study that examined soil, plant, and grasshopper Pb concentrations at increasing distance to a
Zn smelter in China Zhang et al. (2012) found Pb content in all compartments decreased with increasing
distance. Soil Pb ranged from 49.9 to 973.5 mg Pb/kg. Plant Pb concentration ranged from approximately
5 to approximately 65 mg Pb/kg and varied by species (all species serve as a food source for
grasshoppers). Leaf Pb content was greatest in Japanese millet (Echinochloct crusgalli), followed by
Siberian elm (Ulmaspamila) and green foxtail (Setctria viridis). Grasshopper (Locusta migrcttoria
mctnilensis and Acric/a chinensis) Pb content ranged from 1.07 to 46.95 mg Pb/kg (8.83 average). Soil and
plant contamination significantly decreased at 4000 m distance but Pb content in grasshoppers was
significantly higher within only 2000 m to the smelter.
Whole-body Pb content in isopods (Armadillidium granulatum) was positively correlated with Pb
food exposure (100, 500, 1,000 mg Pb/L, dietary values), but concentrations were much lower than food
contamination levels, indicating isopods do not biomagnify Pb (Mazzei et al.. 2013). Simon et al. (2016)
examined soil, leaf litter, and beetle (Carabus violaceiis and I'leroslichiis oblongopunctatus) Pb
concentrations along an urbanization gradient in Hungary. Pb concentration in soils was highest in the
urban locations but not different between rural and suburban locations. There was no difference in Pb
concentration within beetle species across sites but .P. oblongopunctatus (19.6 mg Pb/kg) had higher Pb
concentrations compared with C. violaceiis (not detected). Within/1, oblongopunctatus, Pb concentration
was higher in males compared with females (when pooled across sites). The BAF for .P.
oblongopunctatus was 1.26 in urban environments, 1.48 in suburban environments and 1.37 in rural
environments.
Vinegar fruit flies (Drosophila melanogaster) also display Pb accumulation differences based on
sex. Females had higher Pb accumulation compared with males (Peterson et al.. 2017). Both sexes
exposed to approximately 109 mg Pb/kg (250 (j,M Pb, nominal rearing medium values, concentration in
flies was measured) had higher Pb body concentration (18.44 ng per female versus 7.32 for males)
compared with controls (0.2 ng per male or female), but females had greater concentration values.
Furthermore, exposure of either male or female parent did not lead to generational uptake effects. Pb
loads in unexposed F1 generations with a Pb-exposed parent were no different from those in F1 adults
with control-treated parents. However, in another study by Peterson et al. (2020). when D. melanogaster
were reared in the same conditions but across an increasing gradient of Pb exposure of approximately
109, 217, and 434 mg Pb/kg (250, 500, and 1,000 (j,M Pb nominal rearing medium values, concentration
in flies was measured), they found no effect of sex on Pb accumulation nor a sex-Pb exposure interaction.
Body Pb accumulation did increase with increasing exposure concentrations, but the response was similar
across both sexes. Additional work is needed to determine the effect organism sex has on Pb uptake and
accumulation in D. melanogaster.
Overall, literature since the 2013 Pb ISA adds additional supporting evidence of the importance
of soil variables on uptake and accumulation by soil invertebrates as well as new information on
additional arthropod groups and modifying factors such as season, and possibly, generation. Snails
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typically accumulate Pb at lower concentrations than those found in soil or vegetation, but a higher
concentration of Pb in the hepatopancreas compared with that in the snail foot show uptake via
consumption leads to greater Pb accumulation than uptake through the soil-skin interface. Similarly,
grasshoppers and silkworms readily accumulate Pb but at levels lower than those in both food and soil
contamination. CEC and soil organic content interact with soil Pb concentration on driving uptake by the
common garden snail while pH and Ca content influence uptake and accumulation in earthworms.
Earthworm uptake also depends upon ecotype due to differences in feeding and burrowing behavior. As
discussed in previous assessments, there is an abundance of information examining the effects of
earthworms on Pb mobility and bioavailability due to these feeding and burrowing behaviors. Earthworm
casts, for example, were found to have higher pH and water-extractable OC. Literature since the 2013 Pb
ISA provides new information on the uptake and accumulation of Pb by spiders and butterflies, and
additional information on bees. Generally, Pb concentration is higher in bee bodies compared to honey
and pollen. Two spider genera examined show low accumulation levels in relation to soil contamination,
suggesting spiders do not readily bioaccumulate Pb. Lastly, there appear to be interactions of generation
and sex on Pb uptake by common cutworms and fruit flies, but the results are variable and the overall
effects remain unclear.
11.2.2.4 Uptake and Bioaccumulation in Terrestrial Vertebrates
The 2013 Pb ISA provided evidence of the accumulation of Pb in blood, bones, and a variety of
different tissues in birds and mammals. In studies of birds in the 2013 Pb ISA, the focus was mainly on
ingestion of manmade materials (e.g., Pb shot). In mammals, multiple species were found to accumulate
Pb from contaminated soils as well as from plants grown in contaminated soils. In birds, low dietary Ca2+
concentrations were linked to increased accumulation of Pb in liver, bone, kidney, muscle, and brain
tissues.
New information has become available on the uptake of Pb in terrestrial reptiles and amphibians
since the 2013 Pb ISA. A study of northern pine snakes (Pitnophis melanoleucus melanoleucus) in the
pine barrens of New Jersey found that Pb was accumulated in a wide variety of tissues including liver,
kidney, muscle, skin, heart, as well as in blood, with the highest mean Pb concentration in muscle
(0.393 ±0.131 (ig/g wet weight) (Burger et al.. 2017). The pathway of exposure was not determined in
this study, but the authors suggested that consumption of prey items was the most likely pathway, as pine
snakes are a top predator in their food web. Pb was found to accumulate in the blood of giant toads
{Rhinella marina) captured at industrial, urban, and rural sites in Mexico (Ilizaliturri-Hernandez et al..
2013). Blood Pb levels ranged from 10.8 to 70.6 (ig/dL and were found to increase with increasing soil Pb
levels.
Since the 2013 Pb ISA, new studies have been published that support findings of Pb accumulation
in different mammalian tissues. Tete et al. (2014) and Camizuli et al. (2018) both found evidence of Pb
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accumulation in the kidneys and livers of wood mice (Apodemus sylvaticns). Kidney concentration ranged
from values under the limit of detection to 268.3 |ig/g dry weight, and liver concentrations ranged from
values under the limit of detection to 281.7 (ig/g dry weight. Another study on Pb accumulation in
mammalian tissues evaluated brain tissue from nine mesocarnivore species in Europe (Kalisinska ct al..
2016). Eurasian otters (Lutra Intra), badgers (Meles meles), pine martens (Martes martes), beech martens
(Martes foinci), European polecats (Mustela putoris), red foxes (Viilpes viilpes), feral and ranch American
minks (Neovison vison), raccoons (Procvon lotor), and raccoon dogs (Nyctereutes procvonoides) were all
sampled during this study. Brain tissue Pb was highest in raccoons (0.47 mg/kg dry weight) and lowest in
ranch American minks (0.072 mg/kg dry weight). The study's authors speculated that carrion with
hunting ammunition is likely to be an important source of Pb for omnivores and partial scavengers, while
organic Pb incorporated in the diet and Pb contained in the soil, earthworms, and dusted food may also be
possible sources of exposure.
Studies of bioaccumulation and uptake in birds tend to support information provided in the 2013
Pb ISA and provide additional evidence for Pb accumulation in a variety of different tissues. Soil remains
an important source of Pb exposure in many bird species. French et al. (2017) identified soil consumption
as one of the most common routes of Pb exposure in American woodcocks (Scolopax minor). Woodcocks
use their long bills to probe the soil for earthworms, with their dietary intake comprising as much as 10%
ingested soil, indicating that Pb-contaminated soil may be an important exposure pathway. Additionally,
the consumption of earthworms is another pathway of exposure, as earthworms can bioaccumulate metals
from the soil. Other species with similar feeding habits to woodcock such as American robins (Turdus
migratorins) may be exposed to Pb through these same pathways.
Birds of prey such as bald eagles (Haliaeetus leucocephalus) and California condors (Gymnogyps
californiamis) have also been shown to accumulate Pb in blood and different tissues. A study of bald
eagle nestlings in the western Great Lakes region found blood Pb concentrations ranging from below the
limit of detection to 26.4 (ig/dL wet weight and feather Pb concentrations ranging from below the limit of
detection to 371 |ig/g wet weight (Bruggeman et al.. 2018). The authors speculated that Pb air pollution,
as well as Pb shot and Pb paint may all be sources of exposure. A study of California condors found that
between 1997 and 2010, the annual percentage of condors with blood Pb levels higher than 0.1 |ig/m L
(originally reported as 100 ng/mL) ranged from 50% to 88% (Finkelstein et al.. 2012). However, this
study found that the majority (79%) of condors had blood Pb isotope ratios that were not significantly
different from Pb-based ammunition. This indicates that Pb ammunition is likely the primary source of Pb
exposure in California condors. Behmke et al. (2015) examined bone Pb as a measure of chronic exposure
and Pb in liver as an indicator of more recent exposure in American black vultures (Cora gyps atratus)
and turkey vultures (Cathartes aura) collected in Virginia. Bone Pb was significantly higher than Pb in
liver in both species indicating that Pb in the birds was primarily associated with long-term exposure.
Possible sources of Pb in these long-lived birds based on comparison of Pb isotope ratios in femur bones
and Pb isotope ratios associated with Pb sources included ammunition, coal-fired power plants, leaded
gasoline, and zinc smelting operations.
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In summary, literature since the 2013 Pb ISA (U.S. EPA, 2013) adds support to existing evidence
of Pb accumulation in blood, bones, and a variety of different tissues in terrestrial vertebrates. Pine snakes
accumulated Pb in liver, kidney, muscle, skin, and heart tissue, with the highest concentrations found in
the muscles. In toads, Pb was found to accumulate in blood and increased with increasing soil Pb levels.
New evidence continues to support findings of the accumulation of Pb in tissues from a wide range of
mammalian species. Pb ammunition continues to be a prevalent source of Pb contamination in both
mammals and birds. Consumption of prey species has also been found to be an important route of Pb
exposure especially in species that consume earthworms such as woodcocks and robins.
11.2.2.5 Uptake and Bioaccumulation Through Food Web
In the 2006 Pb AQCD (U.S. EPA. 2006) and the 2013 Pb ISA (U.S. EPA. 2013). various studies
suggested that Pb might be transferred through terrestrial food webs, with lower Pb concentrations
occurring in each successive trophic level. Having data on bioavailable or bioaccessible concentrations of
Pb at every trophic level would lead to more accurate estimates of trophic transfer within food webs.
Since the 2013 Pb ISA (U.S. EPA, 2013), there have been more observational and experimental examples
of gradual attenuation of Pb concentrations with increasing trophic level; however, this depends on Pb
concentration, the presence of other heavy metals, ecosystem, and organism sensitivity to Pb exposure.
Although most of the following studies were conducted in non-U. S. locations or in proximity to point
sources, they further elucidate biotransfer processes for Pb.
Pb was transferred through a soil, nettle, snail food web in Antwerp, Belgium (Boshoff et al.,
2015). In a microcosm field experiment, adult European land snails (Cepaect nemoralis) from an
uncontaminated site were exposed to sites varying in distance from the Umicore Precious Metal Refinery,
a nonferrous smelter in Antwerp, Belgium. The snails were sampled along with nettle (Urticci dioicct), one
of their food sources. Cepcieci nemoralis were placed in microcosms at each site and allowed to feed on
soil, litter, and vegetation for 16 weeks. A subset of snails was collected at weeks 0, 1,2, 4, 8, and 16 for
metal analysis (Pb, arsenic [As], Cd, Cu, Zn, nickel [Ni]) and morphological and physiological biomarker
response (Section 11.2.4.4). Nettle (U. dioicct) samples were collected three times throughout the
experiment for trace metal analysis. Pb concentration in the soil was the only significant factor explaining
Pb concentration in U. diocia. Pb concentrations in the digestive glands of the C. nemoralis varied
spatially and temporally, as there was a statistically significant interaction between site and time. Pb
concentration in the soil was higher than that in U. diocia, while the concentrations of Pb in the digestive
glands of C. nemoralis were similar to or higher than Pb concentrations in U. diocia.
Detoxification may be an important mechanism behind biodilution of Pb with trophic level in the
food web. Silkworms (Bombvx mori) were shown to excrete Pb when fed Pb-exposed mulberry (Mortis
alba) (Zhou et al., 2015). Soils collected from an agricultural field in China were exposed to nominal
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concentrations of 0, 200, 400, or 800 mg Pb/kg via Pb (NCh^. Pb concentrations were also measured in
soil, mulberry, and silkworms. Morns alba was planted in the Pb-spiked soils for 3 months, and the leaves
were collected and fed to fifth instar larvae of B. mori. The available fraction of Pb in the soils, the total
concentration in mulberry leaves, shoots and roots, and B. mori larvae, silk, feces, and adult moth
increased with increasing soil Pb addition in a dose-dependent manner. Roots sequestered the most Pb,
followed by stems, and leaves. The translocation factor was highest for the transfer of Pb from the soil to
the root in the 400 mg Pb/kg treatment, followed by 1.60 in the 800 mg Pb/kg treatment, and 1.13 in the
200 mg Pb/kg treatment. All other translocation factors between the soil and plant (root-soil, stem-soil,
leaves-stem, stem-root, leaf-root, leaves-stem) were below 1.0 or near 1.0 for the control (0 mg Pb/kg).
Across all treatments, the subcellular distribution of Pb in the leaves was greatest in the cell wall,
followed by the soluble fraction, and organelles. Pb treatment did not affect silkworm survival or mean
weight, but increasing Pb treatment negatively affected the silkworm growth rate. Specifically, the body
weight of silkworms was significantly lower at the end of the experiment in the 800 mg Pb/kg treatment
compared to the control and the 200 mg Pb/kg treatment. Pb concentration in the silkworm increased with
increasing treatment. Pb concentration in the feces was the greatest, followed by the concentration in the
peel, the larvae, the silk moths and finally the silk. Metallothionein synthesis increased in B. mori when
fed with Pb-treated leaves. Metallothionein content in the midgut was more sensitive to lower Pb
exposure (200 mg Pb/kg) than metallothionein in the posterior of the silk gland and in the fat body, both
of which increased in the high Pb exposures (400 and 800 mg Pb/kg). These results suggest that B. mori
can detoxify Pb through excretion and homeostasis.
Field studies published since the 2013 Pb ISA (U.S. EPA. 2013) provide additional evidence for
biodilution in terrestrial food webs. Oil rapeseed (Brassica napus) and insects were collected from 35
agricultural sites in Southwest Poland (Orlowski et al.. 2019). These agricultural sites varied in size,
habitat fragmentation, and percent cover by forests and were characterized by percent arable land,
permanent vegetation, linear woody features, dirt or unpaved roads, and wooded areas. Brassica napus
and the insect community (grouped into guilds: pollinators, consumer/herbivores, saprovores, predators,
and parasitoids) were analyzed for Pb and other trace elements. The concentration of Pb in Brassica
napus (mean: approximately 2 mg Pb/kg) was higher than those in all insects examined (range: 0.77 to
2.31 mg Pb/ kg), and Pb concentration generally decreased with increasing trophic level, suggesting Pb is
diluted in this food web. As the size of the field area increased, the Pb concentration in pollinators
decreased, suggesting that even under low Pb levels, larger areas with more diversified landscapes could
reduce Pb body burden for pollinators.
The presence of other heavy metals in the soil, specifically Cd, can affect the uptake and trophic
transfer of Pb. In an agricultural system in Pakistan (Aslam et al.. 2015). alfalfa (Medicago sativa) seeds
were grown in control, Pb (0 mg Pb/kg, 200 mg Pb/kg or 400 mg Pb/kg), Cd (0 mg Cd/kg, 4 mg Cd/kg or
8 mg Cd/kg) or Pb and Cd-enriched soil (200 mg/kb Pb + 4 mg/kg Cd and 400 mg/kb Pb + 8 mg/kb Cd).
Soils were treated with Pb(NC>3)2 and Cd(NC>3)2 salts, resulting in 1.45 ± 0.23 mg/kb Pb (mean ± S.E.) for
control, 112.0 ± 2.43 mg Pb/kg for 200 mg Pb/kg and 237.4 ± 2.79 for 400 mg Pb/kg at the end of the
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experiment for Pb-treated soils. Rabbits (Oryctolagus cuniculus) were placed in chambers and fed with
metal-treated M sativa for 10 days. Soil, M. sativa root and shoot, and O. cuniculus blood and fecal Pb
and Cd concentrations increased with increasing concentrations of metal treatment. Medicago sativa BAF
in the roots increased with increasing Pb concentration and in combined treatments with Cd relative to Pb
exposure alone. Specifically, the Pb BAF associated with the 200 mg Pb/kg treatment was 0.87, while the
BAF resulting from the 400 mg Pb/kg + 8 mg Cd/kg treatment was 0.96. Conversely, Pb contents in the
shoots and leaves ofM. sativa showed higher BAF in the Pb treatments relative to the combined Pb + Cd
treatments. Only a small portion of Pb was transferred to the shoots, as all BAFs were below a threshold
of 1.0. Although not explicitly tested, O. cuniculus blood and feces Pb levels were similar between Pb-
only and Pb + Cd treatments (e.g., fecal concentration in 200 mg Pb/kg treatment: 3.86 ± 0.73 mg Pb/kg
[mean ± S.E.], 200 mg/kb Pb + 4 mg/kg Cd treatment: 2.89 ± 0.67 mg Pb/kg), suggesting Pb uptake and
accumulation is not influenced by the presence of Cd in O. cuniculus. Combined, this study suggests that
although Pb bioaccumulation is higher in M. sativa roots and lower in the shoots in the combined Pb + Cd
treatment relative to Pb exposure alone, it does not affect the uptake of Pb by herbivores such as O.
cuniculus.
Although many observational studies examining BAFs across multiple trophic levels have found
evidence for biodilution of Pb, some studies have observed bioaccumulation. For example, soil samples
(0-15 cm), berseem plants (Trifolium alexandrinum), aphids (Sitobion avenae), grasshopper (Aiolopus
thalassinus) and ladybird beetle larvae (Coccinella septempunctata) were collected from five agricultural
sites in Punjab, Pakistan and analyzed for accumulation of Pb, Cd and Zn. In this study, Pb was not
significantly correlated with any other soil physicochemical variables or metals (percent sand, percent silt,
percent clay, soil OM, CEC, Zn, Cd, or pH). Pb concentrations in the soil were low and similar among all
sites (3.08 ± 0.53 mg Pb/kg, mean ± S.D.). BAFs were greater than 1.0 for Trifolium alexandrinum (BAF
for soil - berseem: 2.26 ± 0.42), Sitobion avenae (BAF for berseem - aphids: 1.40 ± 0.41), Aiolopus
thalassinus (BAF for berseem - grasshoppers: 14.64 ± 3.42). and Coccinella septempunctata (BAF for
aphid - beetle: 2.94 ± 1.31). Overall, this system does exhibit bioaccumulation of Pb, but the
concentrations of Pb in soil were very low. There was no significant correlation between Pb soil
concentration and T. alexandrinum Pb concentration, between S. avenae and T. alexandrinum, between T.
alexandrinum and A. thalassinus and between C. septempunctata and S. avenae.
In summary, Pb generally shows patterns of biodilution through terrestrial food webs; however,
some observational studies have shown bioaccumulation of Pb. Furthermore, the rate at which Pb
biodilutes or accumulates in food webs depends on the presence of cadmium, the sensitivity of the
organism to Pb exposure and ecosystem type.
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11.2.3 Environmental Concentrations of Pb in Terrestrial Biota and
Ecosystems in the United States at Different Locations and Over Time
Studies that present long-term trends of Pb concentrations observed in terrestrial ecosystems are
summarized in this section. National and regional studies that summarize Pb concentrations in soils and
biota on decadal timescales are included.
11.2.3.1 Pb in Soils
Pb concentrations in soils vary across the United States due to a variety of anthropogenic and
natural factors. In general, areas with higher population density and intensity of industrial activity have
higher soil Pb concentrations relative to rural areas. This pattern was observed in the following studies of
national and regional soil Pb concentrations.
A regional survey of forest floor soils sampled in the northeastern United States provides a time
series of Pb concentrations from 1980 to 2011. The region has a large amount of urban and industrial
development associated with high historical anthropogenic Pb emissions. Soils were sampled at 25 sites
in 1980 and sampled again at 16 of those sites in 1990, 2002 and 2011. Sites were located across
northeastern states including Pennsylvania, New York, Connecticut, Massachusetts, Vermont, and New
Hampshire. Across all sites, mean soil Pb concentrations decreased from 151 ± 29 (SE) mg Pb/kg in 1980
to 68 ± 13 (SE) mg Pb/kg in 2011 (Richardson et al.. 2014) (summarized in Table 11-1). The authors
explained the observed reduction in forest floor Pb concentrations by the dilution effect of added organic
material containing less Pb than in older forest floor organic soil as well as by the leaching of Pb from
upper soil horizons into the underlying soil. Isotopic analysis of Pb samples indicated that gasoline was
the dominant source of the measured soil Pb and that it persisted in forest floor soils until at least 2011,
and likely later. In another analysis of the data set of 1980-2011 northeastern U.S. forest floor soils, Pb
concentrations were estimated to decline 2.0 ± 0.3% per year (Richardson et al.. 2015a).
A 2019 survey of peri-urban soil Pb from 54 sampling sites in southern California counties
including Los Angeles, Orange, San Bernardino, and Riverside found that soil Pb was elevated relative to
the southwestern U.S. region, but lower than concentrations found at contaminated sites near point
sources of Pb, with amean of 23.9 ± 13.8 mg Pb/kg (Mackowiak et al.. 2021) (summarized in
Table 11-1). The mean is considerably lower than the forest floor mean observed in the Richardson et al.
(2014) surveys and the results of this study are illustrative of the regional variance in U.S. soil Pb
concentrations. Foliage samples from eight shrub and tree species collocated with soil samples were
collected from the sampling sites of Mackowiak et al. (2021). No correlation was identified between
foliar bioaccumulation and soil Pb concentrations in the study.
Measuring the ratio of Pb concentrations between different soil horizons can provide information
on the relative contribution of anthropogenic Pb to total Pb observed in the soil. In the recent NASGLP
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soil survey of the conterminous United States Smith et al. (2013a) (summarized in Section 11.1.3 and
Table 11-1), samples were collected from multiple soil horizons. Stratified sampling enabled the
comparison of Pb concentrations from bedrock to those in upper-horizon soil. In areas with historic
depositional input of Pb, the concentration of Pb observed in upper-horizon soils was often higher than
that in the bedrock. Figure 11-4C. shows the ratio of A-horizon to C-horizon Pb concentrations mapped in
Woodruff et al. (2015). using inverse-distance weighting methods derived from the NASGLP survey
(Smith et al.. 2013a). This map displays areas with increased concentrations of Pb in A-horizon soils
relative to lower horizons, hinting at the lasting effect of depositional Pb pollution. The mapped ratio of
A-horizon to C-horizon soils from Woodruff et al. (2015) may serve as an indicator for soil in areas
where historical Pb deposition may have a relatively higher effect on people and ecosystems. Patterns of
elevated A- to C-horizon soil Pb concentrations in Figure 11-4C are conspicuous in areas with historical
anthropogenic sources of Pb. This pattern is observed in the northeastern United States, with a historically
high population density and intensity of industrial development. Likewise, mapping highlights former Pb
smelting and mining sites, for instance in areas near smelters in Everett and Tacoma, Washington or the
Doe Run smelter in Herculaneum, Missouri (the last Pb smelter in the United States, which closed in
2013). Areas near mining sites, including near Leadville, Colorado, Cooke City, Montana, and northern
Utah, also have a high ratio of A- to C-horizon Pb. Woodruff et al. (2015) emphasized that no known
natural geological process would otherwise explain elevated A-horizon soils relative to the underlying
layers.
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B. Lead (Pb) - C Horizon
D. Population density (per sq. mile) by county
A. Lead (Pb) - A Horizon
O. Lead (Pb) - Ratio of A Horizon/C Horizon
Source: Woodruff et al. (2015).
Figure 11-4 Maps of Pb sampled from A-horizon (A) and C-horizon (B) soils,
the ratio of Pb observed in A-horizon to C-horizon soils (C) and a
map of U.S. population density (D).
Recent national and regional surveys of soil Pb document the spatial and temporal patterns of
residual pollution from decades of Pb emissions. Data made available from the NASGLP provide the
most comprehensive information on the distribution of Pb across the conterminous United States (Smith
et al.. 2013a). Regional studies of soil Pb provide valuable information on temporal trends and relate
observed soil Pb concentrations to Pb in biota collocated with soil sampling locations. Elevated upper soil
horizon Pb concentrations relative to the underhing soil with greater substratum content observed across
the conterminous U.S. in ( Woodruff et al.. 2015) and over four decades in the northeast in ( dchardson et
al.. 2014) demonstrate the persistence of historical Pb contamination in U.S. soils.
11.2.3.2 Pb in Tree Rings
Dendrochronology can be used to reconstruct historical trends of Pb in air pollution as tree rings
record an annual record of ambient environmental conditions across a tree's lifespan, although radial
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transport of Pb within the tree may reduce the precision of historical Pb concentrations reconstructed from
tree rings. Because trees primarily uptake Pb through their roots, there may be a 10-15-year delay in tree-
ring Pb compared with air Pb concentrations as Pb deposition leaches through the soil and is absorbed by
the tree (U.S. EPA. 2013).
Several studies conducted after the 2013 Pb ISA report temporal trends in Pb as observed in tree
rings, three from Canada and one from Mexico. A study of white spruce trees (Picea glanca) located in
the Northern Athabasca Oil Sands Region of western Canada near oil sands mining operations
reconstructed Pb concentrations from 1878 to 2009. Tree-ring records of Pb concentrations increased
beginning in 1922, peaked in 1968-1973, then decreased until 2009 (Dinis et al.. 2016). In eastern
Canada, a study reconstructed Pb trends from 1880 to 2007 in red spruce (Picea mbens), beech (Fagns
grandifolia), white pine (Pinas strobus), and white cedar (Thuja occidentalis). The beech trees located in
both Montreal and Georgian Bay exhibited a decline in concentrations after a 1970-1985 peak. The
authors attribute the lack of an observed temporal trend in Pb concentrations in white pine to the radial
mobility of Pb within the tree (Douce t et al.. 2012). Another study of tree-ring Pb concentrations in white
cedar in Quebec dated concentrations from 1850 to 2010 and recorded increased concentrations from
1950 to 2000 near a Pb smelter. The increasing trend at a control site further from the smelter was delayed
to 1990-2010. Concentrations across most sites in this study decreased from 2000 to 2010 (Arteau et al..
2020). In contrast to the trends observed in the Canadian studies, a study of Prosopis julifora tree rings
dated from 1903 to 2007 located near a copper smelter in San Luis Potosi, Mexico found increasing Pb
concentrations from 1990 to 2007 (Beramendi-Orosco et al.. 2013).
Although trends in reconstructed Pb concentrations varied across tree species and regions, studies
identified atemporal pattern of Pb that increased after 1850-1900 and, in some cases, peaked in 1970-
1985, then decreased afterward. Tree-ring studies with temporal patterns in exception to this pattern were
conducted near persisting industrial point sources of Pb pollution.
11.2.4 Effects of Pb in Terrestrial Systems
This section focuses on studies of the biological effects of Pb on terrestrial biota published since
the 2013 Pb ISA. First, new information on factors that affect biological sensitivity to Pb is discussed,
followed by subsections on effects on vegetation, microbes, invertebrates, and vertebrates. The biological
effects of Pb in the 2013 Pb ISA and in this appendix are generally presented in increasing order of the
complexity of biological organization, from suborganismal responses (i.e., enzyme activities, changes in
blood variables) to endpoints relevant to the population level and higher (growth, reproduction, and
survival), up to effects on ecological communities and ecosystems.
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11.2.4.1 Effects on Terrestrial Microbes
Several field and laboratory studies have examined the relationship between soil Pb concentration
and microbial community structure and processes. Cell viability of bacteria grown in Pb-contaminated
media was unaffected, and bacteria were able to take up Pb in studies reported in the 1977 AQCD (U.S.
EPA. 1977). Furthermore, in other studies reported in the 1977 AQCD (U.S. EPA. 1977). 1986 AQCD
(U.S. EPA. 1986). and the 2013 Pb ISA (U.S. EPA. 2013). soil Pb concentration was correlated with
decreases in the diversity and function of soil microorganisms. New studies since the 2013 Pb ISA added
a gradient of Pb to the soil and showed negative relationships between Pb concentrations and bacterial
abundance. Most new studies since the 2013 Pb ISA were observational and leveraged natural
environmental gradients of pollutants. In these cases, Pb was not the sole contaminant in the soil,
contributing some uncertainty to their interpretation. Observational field studies showed mixed
associations between soil Pb concentration and microbial abundance and diversity metrics. Additionally,
there has been substantial research on how Pb affects the interactions between microbes and their hosts,
specifically, plants and mycorrhizal associations (Section 11.2.4.2).
Pb contamination slightly affected microbial diversity and significantly affected the abundance of
certain bacteria phyla and genera in an agricultural system (An et al.. 2018). Soil in an agricultural field in
China was supplemented with nominal concentrations of 0, 175, or 350 mg Pb/kg using Pb(NC>3)2 and
permitted to age for 3 months while maintaining soil moisture. After 3 months, soil physicochemical
variables and bacterial community structure were analyzed. Measured available Pb and total Pb
concentration in the soil varied with Pb treatment level (available Pb in the control (mean ± S .D.):
3.97 ± 0.08 mg Pb/kg, 150 mg Pb/kg treatment: 126.6 ± 4.98 and 350 mg Pb/kg treatment:
254.46 ± 7.13). Total and available Pb concentrations were highly correlated. Some soil physicochemical
variables differed between the control and Pb-spiked soils; soil OM was lower in Pb-spiked soils
compared with the control, while soil moisture was the lowest in the 150 mg Pb/kg treatment. Soil pH,
available P, available K, and available N were similar among all treatments. Pb exposure marginally
affected microbial Operational Taxonomic Unit (OTU) richness and diversity, as well as the abundance-
based coverage estimator (ACE), Chao and Shannon's diversity indices were highest in the 175 mg Pb/kg
treatment compared with the control and the 350 mg Pb/kg treatment (statistics not reported, error bars do
overlap). The abundances of certain genera were affected by Pb treatment; Bacillus, Lactobacillus, and
Truepera abundances were negatively correlated with Pb concentration, while Streptococcus and
Arhtorbacter were highest under the low Pb treatment. Bosea and Aquicella increased in abundance with
Pb treatment. Total Pb concentration was correlated with the abundance of Planctomycetes and
Gemmatimonadetes and marginally correlated with Nitrospirae.
Microbial enzyme activity was significantly negatively affected in soils collected from a research
station in northwestern Iran, exposed to nominal concentrations of 0, 100, 200, 300, 400, or 500 mg Pb/kg
using aqueous Pb nitrate and incubated for 2 weeks (Shirzadeh et al.. 2022). After 3, 15, 30, 90 and
180 days, microbial enzyme activities and microbial indices, including acid and alkaline
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phosphomonoesterase, nitrate reductase, urease, soil microbial biomass carbon, soil basal respiration were
characterized. Nominal Pb concentration, incubation time and the interaction between Pb concentration
and incubation time significantly affected all enzyme activities and microbial indices. In general, higher
concentrations of Pb and longer incubation times resulted in a commensurate reduction in enzyme
activities and microbial indices.
The root nodule allocation by the actinobacteria Frankia on Alder (Alnus glutonisa) was
unaffected by Pb treatment, while Frankia microbial respiration was significantly affected by Pb
treatment (Belanger et al.. 2015). The authors suggested that large difference between the maximum
tolerable concentration (MTC), the highest metal concentration when Frankia has 95% of its relative
respiration capacity (<0.01 mM) and the minimum inhibitory concentration (MIC), when under 5% of
relative respiration capacity occurs (10.0 mM), may be due to sequestration or binding of Pb by Frankia,
which has been shown to occur with other heavy metals.
Bacteria and archaeal abundance and diversity have been found to be affected by soil Pb
concentration in several observational studies. Beattie et al. (2018) examined the relationship between Pb
and other soil heavy metals as well as bacterial and archaeal communities in Oklahoma. Picher, an
abandoned mining town, is located near the Picher mine field (PMF), which was declared a U.S. EPA
Superfund Site in 1983 (Tar Creek Superfund Site). Soil samples were analyzed for trace metals and soil
physicochemical properties (Pb, Al, Ar, B, Cd, Cr, cobalt [Co], Cu, Fe, Mg, Mn, molybdenum [Mo], Ni,
K, sodium [Na], tellurium [Te], titanium [Ti], tungsten [W], vanadium [V], Zn, soil pH and soil moisture)
and soil bacterial and archaeal abundance and diversity using 16S rRNA gene copies. Pb soil
concentration was 76.39 ± 1.37 mg Pb/kg (mean ± S. E.) and ranged from 3.0 mg/kb Pb to 1115.2 mg/kg
(Beattie et al.. 2017). Bacterial abundance (16S rRNA gene copies) was found to be negatively correlated
with soil Pb concentration, while archaeal abundance and the bacteria:archaea ratio were not. In addition
to soil Pb concentration, bacterial copy numbers were significantly correlated with Cd, Zn and Mg. Out of
four metals tested (Pb, Al, Cd and Zn), Pb was the only metal to significantly affect microbial diversity.
Shannon-Wiener diversity and Simpson's evenness indices were negatively correlated with Pb
concentration, while the Simpson diversity index was positively correlated, and the Shannon evenness
index was not correlated with Pb concentration. The authors suggested that these conflicting results might
be due to how the indices were calculated or the presence of an outlier. Given that the other metals
analyzed (Al, Cd, Zn) were not correlated with microbial diversity, the authors suggested that the
microbial community had already reached a stable equilibrium with long-term heavy-metal exposure.
Using CCA to determine the relationship of Pb, Cd, Zn and Al with OTU abundance, 1150 OTUs were
found to be significantly correlated with Pb. A total of 2,591 OTUs out of 27,082 were significantly
correlated with one of the four metals (Al, Cd, Pb or Zn), and 60% of these OTUs correlated with two or
more metals while 28% correlated with all four metals. Finally, distance-based linear modeling and
redundancy analysis were used to determine which environmental factors best explained variation in the
soil microbial community. Soil Pb explained 6.96% of the variance in community structure, with only Al
and Zn explaining more (Al = 7.99%, Zn = 7.64%).
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Long-term exposure to Pb and other heavy metals influence microbial community structure, as
heavy-metal-tolerant fungi have been isolated in forested areas in the United States (Torres-Cruz ct al..
2018). Fungi were isolated from soil collected from N-fertilized and unfertilized plots in Duke Forest,
North Carolina. Fungi tolerant to Pb were isolated from the rest of the fungal community by adding
diluted soil to malt extract agar supplemented with antibiotics and Pb stock solutions (100 or 500 ppm
Pb(NC>3)2). Fungal isolates were identified using OTUs and used in phylogenetic analyses and next
generation sequencing was conducted to determine the abundance of heavy-metal-tolerant taxa. The
number of isolated OTUs tolerant to Pb were higher compared with the number of isolates tolerant to
other heavy-metal stock solutions analyzed in this study, including Al, Cr, Fe, Ni, Cu, Cd and Zn, and the
largest number of isolates were obtained from Pb (30% of all isolates) followed by Zn (14% of isolates).
The genera Trichodermct, Penicillium, Umbelopsis, Pochonict, and Saitozvma, all have isolates tolerant to
Pb stress. The most common taxa, Trichodermct and Penicillium, were detected in all metal-enriched
samples, and the authors hypothesized this gives them a competitive advantage across a wide range of
polluted conditions.
Other field studies have found mixed relationships between soil Pb concentration and bacterial
abundance and community structure. For example, Vetrovskv and Baldrian (2015) examined the
relationship between bacteria and actinobacterial biomass and diversity and soil heavy-metal content (Pb,
Cd, Cu, and Zn) across sites ranging in distance from a polymetallic smelter in Pribram, Czech Republic.
Pb soil concentrations ranged from 160.5 ± 3.9 mg Pb/kg (mean ± S.E.) to 1713.5 ± 123.4 mg Pb/kg at
the most contaminated site. Pb concentration in the soil was significantly correlated with Cd, Cu and Zn,
but not oxidizable C, total N content, C/N, and pH. Bacterial biomass, actinobacteria biomass and the
ratio of actinobacteria:bacteria were not significantly correlated with Pb concentration. Finally, the
Shannon-Wiener diversity index increased with increasing heavy-metal contamination.
Although abundance and diversity indices are commonly reported in observational studies
examining the relationship between Pb, other soil metals and microbial communities, some studies have
reported additional effects including average cell wall color development (AWCD) or average carbon
source utilization, microbial growth rate and enzyme activities. These effects can act as surrogates for
microbial activity and diversity. Specifically, Boshoff et al. (2014) used BIOLOG® EcoPlates™ to assess
microbial capacity to metabolize a variety of carbon substrates in two grassland sites that varied in their
distance from an active metal refinery in Antwerp, Belgium. Average carbon utilization AWCD, the
number and variety of utilized substrates (functional richness (R) and the functional diversity (FT)) were
analyzed. Unlike pH, OC, particle size distribution, Cd, Ni and Zn concentration in the soil, Pb
concentration differed significantly between the soils of the two sites, ranging from 147.10 mg Pb/kg to
1,373 mg Pb/kg across all subplots. Additionally, soil moisture, temperature, As and Cu differed between
the two grassland sites. Overall, pseudototal Pb and Cu concentration, which was measured by adding
hydrochloric acid and nitric acid to the samples (as well as As and Cu) was negatively correlated with
AWCD, R' and FT; however, when an analysis of covariance was performed to understand the effect of
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metal pollution on microbial responses, Pb was not a significant factor driving variation for AWCD, R' or
H', while sampling site and As concentration were significant predictors.
In many observational field studies, total Pb soil concentration is often used when analyzing soil
microbial communities; however, some studies attempt to determine bioavailable Pb in addition to total
soil Pb. The relationship between total and bioavailable concentrations of heavy metals (Pb, Zn, Cu, Cd),
soil physicochemical properties (pH, total N, available P, available K and OM) and soil microbial
communities was explored from soil collected near an abandoned ore-dressing plant in Hezhang County,
China (Wang et al.. 2018a). Total soil Pb concentrations ranged from 67.4 ±1.6 mg Pb/kg (mean ± S.D.,
n = 3) to 759.3 ± 11.4 mg Pb/kg, while bioavailable Pb, measured as 0.1 M HCl-extractable Pb (HCl-Pb)
ranged from 33.0 ±1.9 mg Pb/kg to 681.0 ± 33.9 mg/kb Pb. In this study, neither total soil Pb nor HCl-Pb
was correlated with microbial enzyme including fluorescein diacetate hydrolysis activity, an indicator of
soil microbial activity and urease activity. Additionally, Pb was not significantly correlated with any
microcalorimetric parameters examined; however, when bioavailable Pb (HCl-Pb) was used instead of
total Pb, the direction of these trends changed. Pb and HCl-Pb showed mixed relationships with bacterial
abundances. For example, Thiobacilhis, Anaerolineaceae, andXanthobacteraceae abundances were
significantly positively correlated with HCl-Pb, HCl-Cu, and Cu, while uncultured Acidmicrobial.es
showed significant negative correlation with Pb and HCl-Zn.
Previous exposure to pollution in soil may affect the sensitivity of microbial communities in the
rhizosphere to Pb stress (Zhang et al.. 2019b). Ferns (Athyrium wctrdii) were collected from either a site
exposed to mining (mining ecotype or ME) or a reference site (nonmining ecotype [NME]) in Sichuan
Province, China. Collected A wctrdii were then grown in uncontaminated soil for several generations and
subsequently exposed to one of five experimental Pb levels: 0, 200, 400, 600 or 800 mg Pb/kg (aqueous
Pb(NC>3)2). After 50 days, soil Pb concentration, soil respiration, microbial biomass carbon (MBC),
aboveground and belowground biomass, soil physicochemical characteristics (total and available N and P,
pH, and OM), and heavy metals were analyzed. Total and available Pb in the rhizosphere increased
significantly with experimental Pb exposure, while OM, TN, available N, available P, available K, and
pH were similar across all Pb treatments. Total Pb was 9.74 ± 0.11, 210.27 ± 0.41, 412.24 ± 0.60,
607.17 ± 0.65 and 811.74 ± 0.44 mg Pb/kg (mean ± S.D.), and available Pb was 2.15 ± 0.24,
72.23 ± 0.28, 166.30 ± 0.38, 242.94 ± 0.19 and 382.17 ± 0.60 mg Pb/kg, respectively. The rhizosphere of
A. wctrdii ME had significantly higher concentrations (12-4.8 times) of Pb compared with that of the
NME. Microbial activity, characterized through soil respiration and MBC, was reduced under increasing
Pb concentration for both ecotypes; however, the microbial community in the rhizosphere of NME
experienced a greater reduction in MBC when exposed to high Pb treatments (400-800 mg Pb/kg) than
ME plants (NME 28.4%-68.2% versus ME: 21.2%-60.9% less MBC than control). Additionally, the
MBC of soils in the rhizosphere of the NME was significantly lower than that of ME for A. wctrdii
exposed to Pb. Finally, the soil metabolic quotient or soil qC02 increased with increasing Pb exposure;
however, plant ecotype did not affect soil qC02. The authors suggested that in general, the microbial
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community in the rhizosphere of the ME was more adapted to Pb stress than the community in the
rhizosphere of the NME, as soil respiration and MBC are less affected by Pb exposure.
Since the 2013 Pb ISA (U.S. EPA, 2013), additional observational studies, many of which were
natural environmental gradient studies, have linked microbial community structure (e.g., abundance,
diversity) and function (e.g., enzyme activities, respiration rates). Many studies found mixed (negative,
positive, and null) relationships between total or bioavailable Pb soil concentration and the abundance of
bacterial and fungal taxa (Zappelini et al., 2015), diversity (Aleksova et al., 2020; Kerfahi et al., 2020;
Golebiewski et al„ 2014; Tipayno et al„ 2012), microbial C and N (Zcng et al„ 2020), and respiration and
nitrification (Smolders et al., 2015). Unfortunately, it is difficult to disentangle the effects of Pb exposure
on microbial communities from the effects of other soil contaminants using environmental gradients, as
other heavy metals and soil physicochemical properties are significantly correlated with soil Pb
concentration, and many of these factors also influence microbial processes.
11.2.4.2 Effects on Terrestrial Plants and Lichen
In terrestrial plants, Pb is known to induce oxidative stress and impair plant growth, root
elongation, seed germination, transpiration, chlorophyll production, lamellar organization in the
chloroplast, and cell division. However, the extent of these effects varies and depends on the Pb
concentration tested, the duration of exposure, the intensity of plant stress and co-stressors, the stage of
plant development, and the particular organs studied. Plants have developed various mitigations when
exposed to toxic metal exposures including selective metal uptake, excretion, complexation by specific
ligands, and compartmentalization. At sufficiently high Pb exposure, the plant's antioxidant capacity is
exceeded, and peroxidation of lipids and DNA damage follows, eventually leading to accelerated
senescence and potentially, death. In the 2013 Pb ISA, the body of evidence was sufficient to conclude
there are causal relationships between Pb exposure and both plant physiological stress and reduced plant
growth, and inadequate to infer causal relationships between Pb exposure and both plant survival and
plant reproduction (U.S. EPA, 2013).
Previous AQCDs recognized declines in photosynthesis and damage to mitosis as effects of Pb
toxicity in plants (U.S. EPA, 2006, 1986, 1977). The 2013 Pb ISA added additional experimental studies
showing photosynthesis impairment in plants exposed to Pb, and studies of damage to photosystem II due
to alteration of chlorophyll structure, as well as decreases in chlorophyll content in plants, lichens, and
mosses. Recent studies have continued to demonstrate decreases in photosynthetic performance due to Pb
exposure (Alkhatib et al., 2019; Silva et al., 2017a; Rodriguez et al., 2015) as well as documented damage
to chlorophyll structure caused by Pb (Tokarz et al„ 2020; Alkhatib et al., 2019; Rodriguez et al„ 2015).
A substantial amount of evidence of oxidative stress in response to Pb exposure has also been produced
and documented in the 2013 Pb ISA and previous AQCDs. Monocot, dicot, and bryophytic taxa grown in
Pb-contaminated soil or in experimentally spiked soil all responded to increasing exposure with increased
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antioxidant activity. Recent studies continue to confirm increased antioxidant activity in plants in
response to Pb stress (Kaur et al.. 2015; Reis et al.. 2015; Rossato et al.. 2012). as well as the genotoxic
effects of Pb exposure (Silva et al.. 2017b). albeit at concentrations that greatly exceed Pb measured in
soils (Table 11-1). Finally, studies of the effects of Pb on terrestrial plants published since the 2013 Pb
ISA continue to support the previous known findings of declines in plant growth under controlled
exposures of Pb (Muradoglu et al.. 2016; Kaur et al.. 2012; Rossato et al.. 2012).
Although Pb exposure is associated with various responses in plant and lichen species, most
effects seen in terrestrial plants occur at exposures that are generally at higher environmental
concentrations than those outside of the boundaries set for consideration in this ISA (Section 11.1.1).
Additionally, while studies find that exposure to Pb has effects on terrestrial plants that could, depending
on a number of factors, then contribute to community- or ecosystem-scale effects, the exposure methods
typically used make it difficult to compare these effects to what might occur under the uncontrolled
conditions encountered in natural environments. Overall, these experiments demonstrate the effects of Pb
exposure in terrestrial plants and the underlying physiological and biochemical mechanisms, but strong
uncertainties remain regarding the natural concentrations at which these effects would be observed.
One novel area of research is the existence of sex-dependent differences in the effects of Pb in
poplar (Populus spp.) trees. In a study of sexual differences in Populus cathayana exposed to Pb in soil or
applied to the leaves, singly and in combination with drought conditions, Han et al. (2013) found
significantly different effects between male and female trees. Male trees exhibited a greater ability to
bioconcentrate Pb in the root systems, a higher heavy-metal tolerance and photosynthesis plasticity, and
less-damaged cell ultrastructure. When Pb was applied to the leaf alone and in both combined treatments,
there were significant effects on dry mass production, photosynthetic activity, long-term water use
efficiency, potential quantum yield of photosystem II and cellular ultrastructure, and greater effects were
observed in females than in males. Drought worsened Pb stress in both sexes; however, the effects were
larger on female trees. A second study examined sex-dependent responses to Pb stress in the congeneric
Populus deltoides (Xu et al.. 2016b). Pb-induced negative effects on P. deltoides root growth were sex-
related and branch order-specific. Compared with plants in control conditions, Pb decreased total root
length, total surface area, root diameter and biomass, and the effects were significantly greater in female
trees than in males. This agrees with the findings of Han et al. (2013) that female poplar trees exhibit
greater Pb sensitivity. Xu et al. (2016b) found that males of P. deltoides could sequester Pb in the roots of
lower orders and suppress transportation of Pb to high-order roots, which may partially explain the greater
Pb tolerance in males when evaluating tree physiological variables.
Recent studies have also examined the protective effects of certain plant nutrients as well as the
influence of mycorrhizal inoculation on the effects of Pb in terrestrial plants. In a hydroponic experiment
with two different ecotypes of Elshotzia argyi (one from an agricultural site and one from an abandoned
Pb mine in China), plants were exposed to 50 (j,M Pb with normal Zn levels (0.5 (j,M) and high Zn
(20 |iM) for 12 days (Islam et al.. 2011). Application of Pb with normal Zn had negative effects on the
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overall growth and antioxidant capacity of both ecotypes; however, the effects were more pronounced in
agriculturally sourced plants. The addition of high Zn improved the growth and antioxidant capacity of
both ecotypes under Pb stress. Finally, a study using Pb exposures on Torreyci grcindis seedlings (0, 700
and 1400 mg Pb2+/kg) examined the possible protective effects of the addition of 1040 mg/kg Mg2+(Shen
et al„ 2016). The addition of Mg2+ improved the growth of the Pb-stressed seedlings, increased
chlorophyll content, enhanced chloroplast development and improved both the photosynthetic rate and
maximum quantum efficiency in Pb-stressed plants. In addition, Mg2+ addition increased root growth and
oxidative activity and protected the root ultrastructure. These studies showed that some mineral nutrients,
when added beyond the minimal plant requirements, can increase plant tolerance of Pb stress. This is
particularly true of Mg addition. Ling and Hong (2009) hypothesized that Pb2+ may replace either Mg2+ or
Ca2+ in chlorophyll or the oxygen-evolving center, inhibiting photosystem II function through an
alteration of chlorophyll structure.
Mycorrhizal inoculation also appears to protect terrestrial plants from the effects of Pb. One study
examined the effects of AMF (Funneliformis mossecte) on the growth and Pb uptake of Sophora viciifolia
(Xu et al., 2016a). As expected, the AMF altered root growth and architecture (increasing root length,
fork number, tip number, surface area and volume), and these effects are also present under high Pb stress
(1,000 (ig/g). Examining roots under transmission electron microscope and X-ray spectroscopy revealed
that Pb was deposited not only in plant cells but also the cell walls and vacuoles of the AMF intracellular
hyphae, meaning that AMF uptake some of the Pb, alleviating the effects on the plant. Whether the
protective effect of mycorrhizae is species-dependent or not is unknown.
In summary, recent studies have continued to demonstrate various deleterious physiological
effects of Pb exposure, particularly oxidative stress, though uncertainties remain regarding the
environmental concentrations at which these effects would be observed. Additionally, recent studies have
examined the protective effects of mycorrhizae in some plants and of some plant nutrients when added in
excess of plants" minimal requirement. There is still very little evidence addressing the relationship
between Pb exposure and plant survival and reproduction, especially at exposures to concentrations of
interest for this ISA.
11.2.4.3 Effects on Terrestrial Invertebrates
For terrestrial invertebrates, exposure to Pb generally increases mortality, decreases growth, and
can have detrimental effects on behavior as summarized in previous U.S. EPA reviews of this metal. In
studies from the 2006 AQCD, Pb caused antioxidant effects, reductions in survival and growth, as well as
decreased fecundity in soil invertebrates (U.S. EPA, 2006). In the 2013 Pb ISA, there was also evidence
for neurobehavioral aberrations and, in some cases, decreasing fecundity via changes in the endocrine
system (U.S. EPA, 2013). Second-generation effects were reported in some invertebrate species. Recent
literature expands the evidence base for suborganism-level and organism-level endpoints and further
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supports effects on physiological endpoints in additional invertebrate groups, as well as multigenerational
effects of Pb exposure. In addition, recent literature provides new information on the effects of Pb on
organisms not included in the 2013 Pb ISA such as honeybees. Similarly, while soil nematodes are
aquatic organisms—living in the water-filled pore spaces between particles and in water films on soil
particles—they are included in the terrestrial section since they are exposed to soil Pb concentrations.
Accordingly, adherence to aquatic concentration guidelines was not strict when effects were examined in
laboratory conditions.
11.2.4.3.1 Suborganism-Level Response
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a likely causal
relationship between Pb exposure and suborganism physiological level responses in terrestrial
invertebrates (U.S. EPA. 2013). Changes in enzyme activities and oxidative stress markers were reported
in terrestrial invertebrates, including earthworms, snails, and nematodes. Additional studies published
since the 2013 Pb ISA, primarily in earthworms and snails, provide additional supporting evidence for
perturbation of biomarkers of physiological stress associated with Pb exposure.
Available studies in earthworms have assessed a suite of physiological responses including
protein and lipid content following Pb exposure. In field-collected earthworms (Aporrectodea caliginosa)
from metal-polluted soils across northern France, protein content in earthworm was negatively correlated
with easily extractable Pb (CaCh extractable), and stepwise model selection further correlated protein
content positively with soil clay content (Bcaumcllc et al.. 2014). Lipid content was also negatively
correlated with Pb and was positively correlated with silt content. Glycogen was not related to any metal
or soil parameter measured. Total Pb soil concentration varied from 19.6 to 491 mg Pb/kg. It is important
to note that Pb did not occur alone in these soils and is an example of natural pollution conditions. The
authors suggested that energy responses to Pb may be due to demands for mediating oxidative stress
mechanisms or regulation.
Several studies with the earthworm E.fetidct assessed changes in biomarkers of physiological
stress following exposure to Pb. In adult E. fetida exposed via soil (40, 250, 500, 1,000, 2,500 mg Pb/kg,
nominal values, values in organism were measured post-exposure) for 4-weeks followed by a 4-week
recovery period, MDA was higher in Pb-exposed earthworms during both the exposure and recovery
periods (Zaltauskaite et al.. 2020). MDA was positively correlated with soil Pb exposure, and while MDA
concentrations were lower during the recovery period compared with the exposure period, the levels were
still higher than control levels at the end of the recovery period (1.2-1.9 times higher). While MDA levels
did decrease in Pb-exposed worms during the recovery period, the lack of complete recovery of MDA
levels shows worms are not able to recover from Pb-induced oxidative stress within 4 weeks and that
either a longer recovery period is needed or MDA response to Pb has a delayed effect. Juvenile E. fetida
earthworms exposed to Pb had higher levels of MDA, which increased by 25%-54% as soil Pb increased
(40, 250, 500, 1,000, 2,500 mg Pb/kg, nominal values, values in organism were measured post-exposure)
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(Zaltauskaite and Sodiene. 2014). In another study, E. fetida exposed to 5 mg Pb/kg of Pb had lower
protein content than control worms but there was no difference at the 50 and 500 mg Pb/kg exposure
levels (nominal values) (Wu et al.. 2012a). Cellulase activity, however, was higher across all Pb exposure
levels compared with control. DNA damage in coelomocytes (phagocytic leukocytes) was measured by
changes in olive tail moments, tail length, tail DNA content, and tail moment using a comet assay. There
was no effect of Pb on olive tail moments or tail length. Tail moments increased but only in the
50 mg Pb/kg treatment, as did tail DNA contents. The authors concluded since cellulase activity is
involved in the breakdown of cellulose, an increase in cellulase activity suggests Pb may increase E.
fetida s ability to degrade plant matter within the soil profile. Pb exposure at 50 mg Pb/kg appeared to
lead to more DNA damage of coelomocytes but not at 500 mg Pb/kg, indicating more research is needed
to elucidate the effect of Pb exposure on the earthworm immune system via DNA damage, and given that
the exposures were nominal, the putative effects should be quantified with measured exposures in more
complete experiment.
For snails, after 7 days of exposure to Pb via diet, AChE activity in the digestive gland of the
green garden snail (Cantareus apertus) decreased with increasing Pb exposure (nominal dietary exposure
values reported, values in snail tissue measured) (Mleiki et al.. 2015). Activity was 200 (.unol/nm/mg in
control snails, approximately 75 (unol/nm/mg at 25 mg Pb/kg exposure and about 25 (.unol/nm/mg at
2,500 mg Pb/kg Pb exposure. After 60 days of exposure, the activity level was lower across all groups but
followed the same decreasing pattern with increasing exposure. AChE activity in the foot also followed a
similar pattern to the digestive gland, with decreasing activity at day 7 with increasing exposure. After
60 days, differences across treatments were not significant in the foot. Overall, Pb caused a decrease in
AChE activity in both the foot and digestive gland, but the effect was stronger in the short term compared
with the long term. In another snail study, metal concentrations in soil, stinging nettle (Urtica dioica), and
the digestive gland of Cepcieci nemorctlis snails were assessed in relation to the pollution source (metal
smelter in Belgium), with various physiological biomarkers also measured (Boshoff et al.. 2015). Soil Pb
concentrations varied from approximately 50 mg/kg to 1300 mg/kg and generally decreased with
increasing distance from the pollution source. Pb in leaves followed the same general pattern. European
land snails prefer nettle leaves as a food source, and Pb concentrations in the digestive gland followed the
same pattern as those in soil and leaves each week of the experiment, with the pattern becoming more
pronounced over time with far greater concentrations at the pollution source location (orders of magnitude
greater than other sites). Metal concentration in plants was positively correlated with soil concentrations,
and concentrations in the snail digestive gland were positively correlated with plant concentrations.
Protein, glycogen, GST, and total energy levels measured within the digestive gland showed no clear
pattern in relation to Pb and instead depended on interactions between the specific site, exposure time,
and different heavy metals. There were also no correlational changes in shell morphology.
Physiological stress response linked to Pb exposure was reported in a few additional terrestrial
invertebrates. Overall, gut enzyme activities, with the exception of alpha-glucosidase, were higher in
honeybees (A. mellifera) within urban-located hives in Nigeria compared with wild beehives.
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Carbohydrases (amylase and cellulase) were higher than lipase and proteinase across both nesting
habitats. However, there was no difference in Pb concentration in bees between habitats, and differences
in enzyme activities showed no direct correlation to Pb specifically (Lawal et al.. 2014). In another study,
honeybees in a laboratory setting were fed a sucrose solution with Pb concentrations of 10, 1, 0.1, and
0 mg Pb/L over a 48-hour period. GST enzyme activity and gene expression were examined, along with
AChE activity. No effect of Pb was observed at any exposure concentration on GST activity or gene
expression after 48 hours. AChE activity was lower at 0.1 mg Pb/L and higher at 10 mg Pb/L
concentrations (Nikolic et al.. 2019). In atrophic study examining Pb uptake by mulberry trees (M alba)
and subsequent transfer to silkworms (B. mori), Pb content in silkworms and silkworm excretions (feces
and silk) increased with increasing Pb treatment (0, 200, 400, and 800 mg Pb/kg soil treatments, nominal
values) across lifestages (larvae and moth). Additionally, metallothionein was higher in the midgut in all
Pb treatments compared with control larvae and was higher in the 800 mg/kg treatment compared with the
200 and 400 mg Pb/kg treatments. Metallothionein was also higher in silk-glands and body fat in the 400
and 800 mg Pb/kg treatments (Zhou et al.. 2015).
11.2.4.3.2 Organism-Level Response
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a likely causal
relationship between Pb exposure and neurobehavioral responses in terrestrial invertebrates (U.S. EPA.
2013) (and see Table 11-2 of this appendix). Evidence was primarily from feeding studies in snails and
altered behaviors in nematodes (Caenorhabditis elegans). Several new studies have assessed behavior
modification following Pb exposure in soil organisms and flying insects; most were conducted at nominal
dietary Pb concentrations, whereby known nominal amounts of Pb were fed to the organisms as a method
of achieving a gradient of tissue concentrations that was in turn measured.
Additional studies in nematodes lend further support to Pb neurotoxicity in these organisms. In a
behavioral food preference and food-finding lab study using agar plates, nematodes (C. elegans) avoided
contaminated food and chose uncontaminated food spots at 1 mg Pb/L, but at 129 mg Pb/L (50% lethal
concentration; LC50), Pb contamination interfered with food-finding ability, and there was no difference
in movement toward either contaminated or uncontaminated food (Monteiro et al.. 2014). Another study
using C. elegans found that feeding activity decreased as Pb concentration increased. EC50 for feeding
behavior was approximately 15 mg Pb/L (54 (iM). Pb also increased damage to the dopaminergic neurons
(Tang et al.. 2019). The study also examined the effects of Cd and Mn, the effects when Pb was mixed
with either metal, or the effects of a treatment containing all three metals. The effects on C. elegans
feeding behavior were greater than the additive effect in binary Pb mixtures at fa < 0.85 (fraction of
organism system affected) but less-than-additive at fa >0.9. The ternary combination had greater-than-
additive effects at fa < 0.75 and less-than-additive effects at fa > 0.8.
New studies in honeybees suggest Pb exposure alters feeding and foraging behaviors. Soil Pb
contamination (approximately 47.3 mg Pb/kg) did not change the number of honeybee, bumblebee, or
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megachilid visits to sunflowers but soil contamination did change the foraging behavior of bees (Sivakoff
and Gardiner. 2017). Bumblebees visited uncontaminated grown sunflowers 5.4 times, honeybees 3.7
times and megachilidae 3.6 times longer than sunflowers grown in contaminated soils. Structural equation
modeling analysis shows a direct negative effect of Pb soil contamination on bee visit duration for
bumblebees and honeybees but direct effects of floral traits or indirect effects of Pb on floral traits were
not significant, suggesting Pb contamination directly explains bee visit duration when floral traits are held
constant. In a behavioral lab experiment, A. mellifera were exposed to a range of Pb concentrations (0.07,
0.66, 6.61, 661 mg Pb/kg, dietary values) in a sucrose solution to examine the effect of Pb contamination
on feeding behavior. Only at the highest Pb concentration did bees reduce sucrose solution intake. By
measuring neuron response to sucrose in antennal gustatory sensilla, the authors determined this response
was due not to detection of the Pb but rather due to a decrease in sucrose perception when Pb was added
to the solution^ Furthermore, bees readily ingested the Pb-contaminated solution within a range of 0.075
to 0.75 mg Pb/kg, which the authors reported as comparable to concentrations found in flowers (1.1 to
1.735 mg Pb/kg) (Monchanin et al.. 2022). In another behavioral honeybee experiment, the effects of Pb
(0.07 and 0.66 mg Pb/kg, dietary values) on bee cognitive flexibility were tested. Bees exposed to
0.66 mg Pb/kg contaminated food over 70 days showed less flexibility in response to changing flower
rewards. This response was positively correlated with bee body Pb concentration. Furthermore, higher Pb
exposure during the larval state correlated with lower body weight and head size (Monchanin et al..
2021).
A behavioral experiment examined whether there was a difference in foraging behavior between
cabbage white butterflies (Pieris rcipcie) reared on a Pb-contaminated diet versus those raised on an
uncontaminated diet (Philips et al.. 2017). Larvae were fed either a 4 mg Pb/kg (dietary values) or control
(approximately 0.17 mg Pb/kg) diet. Behavioral testing following Pb exposure involved yellow sponges
soaked in honey (rewarding) or water-soaked blue sponges (nonrewarding). Butterflies reared on Pb as
larvae were more likely as adults to interact with sponges (approximately twice as many adults interacted
with the sponges compared with control-reared butterflies). Of the butterflies that did interact with the
sponges, there was no difference between treatment groups in the proportion that completed five
consecutive landings on the rewarding sponge. There was also no difference in the duration it took for
butterflies to complete the test (time taken to land five times in a row on yellow sponges). The authors
suggested this species may already have adapted to low levels of Pb in their diets because brassicas
(natural food source for larvae of P. rcipcie) mature quickly and are often found in disturbed locations
where Pb may be present. Therefore, the 4 mg Pb/kg concentration may not have been high enough to
induce a different response between treatments in the laboratory-exposed butterflies.
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a likely causal
relationship between Pb exposure and growth endpoints in terrestrial invertebrates (U.S. EPA. 2013) (see
Table 11-2 of this appendix). Evidence in the 2013 Pb ISA was primarily from concentration-dependent
inhibition of growth in earthworms raised in Pb-amended soil, and, to a more limited extent, for reduced
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growth in snails (dietary studies) and nematodes. New evidence continues to show growth related effects
in invertebrate soil organisms.
Additional studies in earthworms since the 2013 Pb ISA continue to support findings of Pb on
growth. Zaltauskaite et al. (2020) examined the effects of Pb exposure on earthworm (E. fetidci) weight,
growth, and recovery postexposure. During 4 weeks of soil exposure (40, 250, 500, 1,000,
2,500 mg Pb/kg, nominal values, values in organism were measured post-exposure), no effect on weight
loss was found, but Pb decreased growth rate with a difference of 15.8%—40% lower fresh weight
compared with control worms. Following exposure, earthworms were given a 4-week recovery period
with no Pb exposure. While earthworms recovered some weight, they did not reach equal weights
compared with non-exposure worms (11%—17.6% lower than control at end of recovery period). Fresh
weight was negatively correlated with increasing Pb soil concentration during both the exposure and
recovery periods. Growth and recovery rate varied with concentration, with earthworms exposed to
40 mg Pb/kg having the greatest growth rate compared with other Pb concentrations. Earthworms grew
slower during the recovery period compared with the exposure period except for those exposed to
2,500 mg/kg, which showed equal growth rates during exposure and recovery. MDA was also positively
correlated with Pb levels. During recovery, MDA concentrations were lower but did not reach the same
levels as control worms. Weight response to Pb exposure and recovery suggests Pb inhibits earthworm
growth and may have a short-term legacy or lag effect as recovery did not reach 100% within the same
time frame. Increased MDA concentration is indicative of oxidative stress, which may explain the
reduced growth since MDA concentrations were still comparatively high after the recovery period.
Another earthworm study by Zaltauskaite and Sodiene (2014) examined juvenile earthworm growth and
time to maturation across nominal soil Pb concentrations of 40, 250, 500, 1,000, 2,500 mg Pb/kg, values
in organism were measured post-exposure. There was no overall effect on weight loss, but juveniles
exposed to Pb were smaller than control worms. The EC50 for juvenile growth increased with increasing
time of exposure—at 3 weeks, the EC50 was approximately 100 mg Pb/kg but after 14 weeks the EC50 for
reduced weight was 179 mg Pb/kg. The time of maximum growth in the 40 mg Pb/kg exposure group was
during the 8-10-week period, while maximum growth was delayed in higher Pb treatments. Pb
significantly lengthened the time to sexual maturation. The minimum time to maturation was 9 weeks for
the control and Pb treatment groups, and the minimum weight at this development point was 0.182 g in
the 40 mg Pb/kg treatment group. Since increasing Pb concentrations reduced the growth rate, the time
needed to reach the minimum maturation size would increase with increasing Pb; therefore, the time
needed at 250 mg Pb/kg would be 16 weeks. The total number of earthworms that reached maturity by the
end of the experiment was negatively correlated with Pb concentrations, with only 5%-7% of worms
reaching maturity in the 250 mg Pb/kg treatment group.
Adding to the evidence for growth effects in snails from the 2013 Pb ISA, studies on green
garden snail (Cantareus apertus) bioaccumulation and growth in response to increasing Pb dietary
concentrations (25, 100, and 2,500 mg Pb/kg, dietary values) over 1 week and 8 weeks of exposure found
the wet weight of snails increased with time across all Pb treatments, and the effect was dose-dependent
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in Pb-treated snails (Mlciki et al.. 2016). The weight of snails was significantly lower than the weight of
control snails by week 2 in the high Pb-treatment group, by week 3 for medium Pb-treatment snails and
by week 7 for snails in the low Pb-treatment group. The cumulative growth rate followed a similar pattern
but was lower by week 1 for the high Pb-treatment snails, by week 3 for medium treatment and by week 7
for low Pb treatment. Overall, dietary Pb decreased growth in green garden snails, with a lowest observed
effect concentration (LOEC) of approximately 25 mg Pb/kg food within several weeks. A trophic snail
study found soil Pb levels varied from approximately 6 mg Pb/kg to 52 mg Pb/kg across a gradient of
polluted sites in Romania (Nica et al.. 2012). Shell height was negatively correlated with Pb in nettle
leaves (food source), and relative shell height was positively correlated with snail hepatopancreas Pb
levels. Pb in soil was also correlated with other metals (Zn and Cd). Heavy metals are known to
accumulate in snail shells and can often lead to changes in shell size and geometry.
The growth effects of Pb reported for earthworms, snails and nematodes are augmented by
studies in a few additional terrestrial invertebrates. In a generational study with tobacco cutworms
(Spodopterci litiira) reared on artificial diets with increasing Pb concentration, both Pb and generation
effects were observed on relative growth rate, pupation rate, and eclosion rate (Shu et al.. 2015). First-
generation pupae experienced no effects of Pb stress on pupation rate or relative growth rate. Eclosion
rates did decrease in the 100 and 500 mg Pb/kg treatments groups (dietary values) (eclosion rates were
51.48% and 28.89%, compared with approximately 70% for all other treatments). Fifth generation larvae
showed significantly lower eclosion and pupation rates at 25 and 50 mg Pb/kg compared with
12.5 mg Pb/kg and control treatments. The relative growth rate of fifth generation pupae declined as well
for the 25 and 50 mg Pb/kg treatments. Differences between generations occurred at the 50 mg Pb/kg
treatment, with 50 mg Pb/kg having stronger negative effects in the fifth generation compared with the
first. There was no effect of Pb (4 mg Pb/kg) on cabbage white butterfly (Pieris rapae) development time
or body size regardless of Pb concentration or butterfly sex (Philips et al.. 2017). Kenig et al. (2013)
reared fruit flies (Drosophila siibobsciira) in the lab for eight generations at low and high Pb exposure
(10 (ig Pb/mL and 100 |ig Pb/mL, dietary values) from two wild-caught populations with a difference in
Pb exposure history (298.6 mg Pb/kg and 25.7 mg Pb/kg soil). Flies from the population originally
collected from the site with high pollution levels exhibited a decrease in development time over
generations reared at control (no Pb) lab conditions, a decrease in development time when reared at low
Pb-exposure lab conditions and an increase when reared at high Pb-exposure conditions. Flies from the
low historic contamination site exhibited an increase in development time at control conditions, a
decrease at low Pb exposure, and a decrease at high exposure. Across all levels of Pb exposure in the lab,
there were population, generation, and population x generation effects on fruit fly development time.
Overall, the flies from the high Pb-exposure contamination group had faster development time across
both lab exposure Pb concentrations compared with the low historic contamination population responses.
The authors suggest this response in development time in the high historic exposure population may be an
ancestral adaptation response to allow for growth and reproduction to occur before Pb toxicity occurs.
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In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a causal relationship
between Pb exposure and reproduction in terrestrial invertebrates (U.S. EPA. 2013) (see Table 11-2 of
this appendix). Reproduction endpoints examined in the 2013 Pb ISA included brood size and hatching
success. Additional studies in soil invertebrates published since the 2013 Pb ISA continue to report Pb
effects on reproduction and development, adding to the evidence base for this endpoint.
Pb exposure (40, 250, 500, 1,000, 2,500 mg Pb/kg, nominal values, concentration in tissue
measured) significantly lengthened time to sexual maturation for juvenile E.fetidct earthworms
(Zaltauskaite and Sodiene. 2014). The minimum time to maturation was 9 weeks for the control and Pb
treatment groups, and the minimum weight at this development point was 0.182 g in the 40 mg Pb/kg
treatment group. Since increasing Pb concentrations reduced growth rate, the time needed to reach the
minimum maturation size would increase with increasing Pb; therefore, the time needed at 250 mg/kg
would be 16 weeks. The total number of earthworms that reached maturity by the end of the experiment
was negatively correlated with Pb concentrations, with only 5%-7% of worms reaching maturity in the
250 mg/kg treatment group. In addition, cocoons were only found at the lowest treatment of 40 mg Pb/kg,
and the number of cocoons was less than half of the number of cocoons produced in control soils.
In a multigeneration vinegar fruit fly (D. melcmogaster) study, females that were reared under no
Pb conditions preferentially mated with control males (60% of the time) over males reared in Pb
conditions (108 mg Pb/kg) (Peterson et al.. 2017). In the same study, Pb-reared females preferentially
mated with Pb-reared males over control males (65% of the time). Second-generation females did not
show a significant preference for either second-generation male group (Pb-reared mother or control-
reared mother). Males across treatments showed no mate preference, and second-generation male body Pb
content was not related to parental Pb content. Despite the behavioral response of females in mate
preference, a principal component analysis of male and female pheromones showed no significant
difference between either male or female treatment groups. Furthermore, there was no difference in
multiple male courtship song variables. While the mechanisms for mate preference remain unclear, there
appears to be no generational effect on fitness. There was no difference between Pb treatments in the
parental generation on either parental or second-generation responses in dry body weight, fecundity, or
time to reach either 50% or 80% mortality. Pb accumulates in fruit fly bodies and this accumulation
appears to influence female but not male mate choice but does not lead to any differences in ability,
success, or fecundity of the flies or their offspring. Another study with D. melcmogaster observed that
vinegar fruit flies accumulate Pb linearly with Pb exposure concentration and that the number of eggs laid
on Pb-treated media varied with Pb treatment (Peterson et al.. 2020). Control-reared females laid fewer
eggs on Pb-contaminated media than Pb-reared females at both approximately 109 and 217 mg Pb/kg
(250 and 500 (iM, PbAc). However, females reared on the highest Pb treatment of approximately
434 mg Pb/kg (1,000 (j,M) laid fewer eggs than the other Pb treatment females. These results suggested
females reared in a Pb-free environment avoid laying eggs in Pb-contaminated areas whereas females
raised in a Pb-contaminated environment did not show this preference for egg site. The authors suggested
this may be due to a loss of this specific avoidance behavior due to developmental exposure or possibly
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due to changes in microbial composition. The microbial composition influences oviposition site selection,
with females choosing a site with a composition more similar to the one in which they grew. Pb acetate
was used as the source of Pb contamination in this study. Pb acetate may directly change the microbial
community, which could also explain why Pb-reared females did not discriminate in laying their eggs in a
Pb-contaminated site.
Kenig et al. (2013) isolated Drosophila subobscura adults from wild populations collected at two
sites with different Pb contamination histories (high pollution site 298.6 mg Pb/kg soil average and low
pollution site of 25.7 mg Pb/kg). Gravid females from both populations were used to establish separate
population breeding lines. Flies were then reared for multiple generations on either a control substrate (no
Pb contamination), a low Pb contamination substrate (10 (ig Pb/mL, dietary values) and a higher Pb
contamination substrate (100 |ig Pb/mL, dietary values). Reproduction response variables were measured
at the F2, F5, and F8 generations for each of the two population lines. Both populations reared under
control conditions in the laboratory across eight generations exhibited an increase in the number of eggs
laid between the F2 and F5 generation. This was followed by a decrease in egg production by the
F8 generation but only for the population with a lower historic Pb exposure. Under low Pb-exposure lab
conditions, both populations showed the same pattern of increasing number of eggs from F2 to F5
followed by a decrease in production to F8, though this pattern was less pronounced for the low historic
exposure population. Under high exposure conditions, both populations saw egg production decrease by
the F8 generation. Egg viability for the high historic exposure population decreased from F2 to F5/F8
under control conditions, and the low exposure population saw an increase from F2 to F5 followed by a
decrease to F2 viability levels by generation F8. Under low exposure conditions, both populations
followed the same pattern they showed under control conditions. Under high Pb lab conditions, neither
population showed a change in egg viability across generations but the egg viability of the population
from low historic exposure conditions had overall lower egg viability than the population that experienced
historically high exposure. Individuals from the historic high exposure showed higher viability and
fecundity when exposed to higher Pb concentrations in all generations compared with those from the
historically low exposure population, exhibiting higher tolerance to heavy-metal exposure.
Mazzei et al. (2013) examined isopod Armadillidium granulatum reproductive response to metal
contamination of food. According to the authors, isopod heavy-metal concentration factors vary widely
across species as does their breeding patterns. In this study, Pb concentration (100, 500, 1,000 mg Pb/kg,
dietary values) in food led to an alteration of reproductive patterns in A. granulatum. Increasing
concentrations led to a delayed onset in breeding season while also reducing the duration of the season.
Breeding season onset did not differ between control and 100 mg Pb/kg treatments. Breeding season
occurred 1 week later in the 500 mg/kg treatment group and 6 weeks later in the 1,000 mg Pb/kg
treatment group. The length of the breeding season decreased from 79 days (control) to 59
(500 mg Pb/kg) and 46 (1,000 mg Pb/kg) days. There was no effect of Pb on incubation period
(approximately 23 days), and the percent gravid rate of females increased from 97.2% (control) and
95.8% (100 mg Pb/kg) to 100% for higher Pb treatments. However, while gravid rate increased, brood
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number declined (from 1.22 to 1). Lastly, the number of juveniles for each brood increased with
500 mg Pb/kg treatment. Overall, contamination at 100 mg Pb/kg did not influence any reproductive
endpoint examined for A. gramdatum but higher levels led to changes in breeding seasonality and the
number of juveniles.
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a causal relationship
between Pb exposure and survival in terrestrial invertebrates (U.S. EPA. 2013) (see Table 11-2 of this
appendix). Additional evidence continues to show Pb effects on mortality in some terrestrial
invertebrates, while others appear to be unaffected. In a laboratory study examining green garden snail
(Cantareus apertus) response to increasing Pb dietary concentrations (25, 100, 2,500 mg Pb/kg, dietary
values) over a period of 1 to 8 weeks, cumulative mortality was greater in all Pb treatments than the
control after 6 weeks of exposure, with the high treatment having significantly greater mortality after
1 week (Mleiki et al.. 2016). At the end of the experiment, cumulative mortality was below 30% for all
treatments. An observational study of C. apertus exposed to multiple metal-polluted soils with Pb
concentrations ranging from 28.1 to 4574 mg Pb/kg found only 6.5% of snails died after 28 days of
exposure (Pauget et al.. 2013b). Studies in the 2006 Pb AQCD found earthworm LC50 for 14 and 28-day
exposure fell within a range of 2,400-5,800 mg Pb/kg. A study reported in the 2013 Pb ISA evaluated E.
fetida earthworms exposed to field-collected soils with Pb concentrations up to 390 mg Pb/kg and found
no effect on earthworm survival (Delistratv and Yokel. 2014). In support, juvenile E. fetida earthworms
exposed to a range of Pb concentrations (40, 250, 500, 1,000, 2,500 mg/kg, nominal values, concentration
in tissue measured) over 14 weeks found mortality increased in the 500, 1,000 and 2,500 mg Pb/kg
treatments, with mortality reaching 90% in the highest treatment (Zaltauskaite and Sodiene. 2014).
Juvenile mortality increased with the time of exposure in these treatment groups, with an LC50 of
911 mg Pb/kg for 14 weeks of Pb exposure. Juvenile mortality did reach 10% by week 3 for the 40 and
250 mg Pb/kg treatments but did not increase any further overtime. However, in another earthworm
exposure experiment using adults of E. fetida, across only 4 weeks of exposure (40, 250, 500, 1,000,
2,500 mg Pb/kg, nominal values , concentration in tissue measured), there was no significant effect on
survival (Zaltauskaite et al.. 2020). In cabbage white butterflies (P. rapae) raised from eggs from wild-
caught females, no effect on survival was observed in a laboratory study with a diet of 4 mg Pb/kg
(Philips et al.. 2017).
In terrestrial invertebrates, literature since the 2013 Pb ISA provides additional support on the
effects of Pb exposure on organismal and suborganismal responses including a decrease in survival and
reduced growth and fecundity. Recently published studies on physiological responses to Pb included
decreases in protein and lipid content and increases in MDA in earthworms. AChE activity decreased in
response to Pb in snails and honeybees while protein, glycogen, other enzymes, and GST responses were
variable depending on modifying site factors or species examined. There are several new studies
quantifying behavioral changes to Pb exposure in bees. Soil Pb contamination altered foraging behavior,
and at high levels (above 600 mg Pb/kg), also altered sucrose intake. However, at low concentrations
(0.66 mg Pb/kg), honeybees showed lower flexibility in response to changing flower rewards, suggesting
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Pb may lead to lower nectar and pollen supply and subsequently slower colony development or winter
survival. New literature on growth endpoints suggests Pb can have lasting effects even postexposure on
earthworms. Growth, eclosion, and pupation rates of the common cutworm were all lower under Pb
exposure, and fruit fly development time increased within eight generations in populations with historic
Pb pollution exposure. In addition to previously assessed endpoints of Pb on brood size and hatching
success, new literature shows Pb exposure slows time to maturation in earthworms, delays onset to and
duration of breeding season in isopods and influences mate selection in fruit flies. While the literature
since the 2013 Pb ISA has primarily provided additional support on previously examined organisms and
endpoints, there has been new information on new organisms as well as on modifying factors on organism
response including habitat, exposure history, seasonality, and duration of effects.
11.2.4.4 Effects on Terrestrial Vertebrates
In observational and experimental studies, commonly observed effects of Pb on terrestrial
vertebrates include decreased survival, reproduction, and growth, as well as effects on development and
behavior (U.S. EPA, 2006). The 2013 Pb ISA (U.S. EPA, 2013) also provided evidence for Pb effects on
hormones and other biochemical variables (U.S. EPA, 2013). Recent studies provide additional support to
suborganism-level and organism-level endpoints and expand on the effects on hematological and
physiological endpoints.
11.2.4.4.1 Suborganism-Level Response
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a causal relationship
between Pb exposure and hematological effects in terrestrial vertebrates (U.S. EPA, 2013) (see
Table 11-2 of this appendix). Since the 2013 Pb ISA, numerous new studies have continued to support the
connection between Pb exposure and hematological effects. The relationship between Pb concentrations
and aminolevulinic acid dehydratase (ALAD) activity has been explored in the literature, across a broad
assortment of different vertebrate species including songbirds (Beyer et al., 2013), house sparrows (Cidet
al., 2018), Japanese quail (Beyer et al„ 2014), griffon vultures (Espin et al„ 2015), eagle owls (Espinet
al., 2015), common ravens (Herring et al„ 2018), turkey vultures (Herring et al„ 2018), Canada geese
(van der Merwe et al., 2011), mallards (Binkowski and Sawicka-Kapusta, 2015), coots (Binkowski and
Sawicka-Kapusta, 2015), giant toads (Ilizaliturri-Hernandez et al„ 2013), cattle (Rodriguez-Estival et al.,
2012), and sheep (Rodriguez-Estival et al„ 2012).
Beyer et al. (2013) investigated blood, liver, and kidney concentrations of Zn, Cu, Pb, and Cd and
ALAD activity in northern cardinals (Cardinalis cardinalis) and American robins (Turdus migrcttorius)
living in Pb-contaminated mining sites in southeast Missouri. Birds from contaminated locations had
ALAD activity levels that were decreased by between 5 8% and 82% compared with those from
noncontaminated locations. Another field study that examined the relationship between Pb and ALAD
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activity found similar results in griffon vultures (Gyps fulviis) and eagle owls (Bubo bubo) (Espi'n ct al..
2015). Blood samples were taken from birds near an industrial area (electric power plants, explosives, and
ship-building factories) and a historic Pb-Zn mine. The study found a significant negative relationship
between blood Pb levels and ALAD activity in griffon vultures and in eagle owls, with ALAD inhibition
of up to 94% and 79%, respectively.
Herring et al. (2018) examined the effects of Pb exposure on ALAD activity in two species of
free-living scavengers in the Pacific Northwest: common ravens (Corvus corax) and turkey vultures
(Cathartes aura). The authors speculated that environmental Pb exposure in these species was most likely
associated with a variety of sources including hunting, Pb-based paint, soil, and sediment Pb, and mining
and smelting activities. Both species exhibited decreased ALAD activity (mean = 5.9 ± 1.4 SE) in birds
with blood Pb concentrations greater than 0.2 (ig/g (the subclinical toxicity benchmark) when compared
with birds with blood Pb concentrations below this benchmark (mean = 9.9 ± 0.6 SE).
Binkowski and Sawicka-Kapusta (2015)is another field study that examined the relationship
between blood Pb levels and ALAD activity in free-living birds published since the 2013 Pb ISA. This
study investigated free-living mallards (Anas platyrhynchos) and Eurasian coots (Fiilica atra) in Poland.
In both species, there was a significant negative correlation between Pb concentrations in blood and
ALAD activity. The authors suggested that Pb exposure mainly occurred through Pb shot, van der Merwe
et al. (2011) also found evidence of a relationship between Pb concentrations and ALAD inhibition in
waterfowl. Geese from the tri-state mining district of Kansas, Oklahoma, and Missouri and multiple
different metal concentrations were measured (silver [Ag], As, barium [Ba], Cd, Co, Cr, Cu, Fe, Mg, Mn,
Mo, Ni, Pb, Se, Ti, V, Zn). This study found that ALAD activity was inversely correlated with tissue Pb
concentrations in all tissue except muscle.
Multiple laboratory studies have examined this relationship. Cid et al. (2018) exposed house
sparrows (Passer domesticus) to sublethal oral doses of Pb acetate solution (1.3, 3.5, 5.5, 7.0, 14.0 |ig/g
animal/day) for 5 days. This resulted in a gradual decrease in ALAD activity between 3.5 and 7.0 |ig Pb/g
animal/day, with the 7.0 and 14.0 |ig Pb/g animal/day doses producing greater a-ALAD activity inhibition
(82% less activity than control group). This study also examined the effects of Pb exposure in drinking
water for 15 or 30 days. Inhibition of ALAD activity was similar between the two groups, with an
approximately 35% decrease when comparing the mean value of both treatment groups and the controls.
Bever et al. (2014) studied the effect of Pb-contaminated soil on captive Japanese quail (Coturnix
japonica) to examine the relationship between Pb exposure and hematological effects and to determine
benchmark doses associated with different percentages of ALAD reduction. Quail were fed experimental
diets containing 0% to 12% contaminated soil by weight (0.12 to 382 mg Pb/kg, dry weight) for 6 weeks.
All quail groups exposed to Pb-contaminated soil had a significantly lower mean ALAD activity than the
control group. ALAD activity also decreased with increasing dosage, with control quail having the
highest amount of activity and the 12% contaminated soil group having the lowest. The benchmark doses
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of Pb associated with a 50% reduction in ALAD activity were 0.62 mg Pb/kg in the blood, dry weight,
and 27 mg Pb/kg in the diet.
Although there is limited new evidence on the effects of Pb on ALAD activity in other terrestrial
vertebrates since the 2013 Pb ISA (U.S. EPA. 2013). two nonbird studies examined this relationship.
Rodn'gucz-Estival et al. (2012). investigated this relationship in both cattle and sheep from livestock
farms in Spain. Blood Pb level was found to be negatively correlated with ALAD reaction ratio in both
cattle and sheep. Blood Pb level also had a negative effect on ALAD activity. Ilizaliturri-Hernandez et al.
(2013) examined the relationship between blood Pb levels and ALAD inhibition in giant toads (Rhinella
marina) in Veracruz, Mexico. Blood Pb levels ranged from 10.8 to 70.6 (ig/dL and were significantly
higher in industrial sites. Toads at industrial sites also had a 78% decrease in ALAD activity when
compared with those at rural sites. Examining the relationship between blood Pb levels and ALAD, a
strong inverse relationship was identified. The authors stated that Pb exposure was most likely from
pollution released into the air and water by chemical and petrochemical companies in the area.
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a likely causal
relationship between Pb exposure and physiological stress for terrestrial vertebrates(U.S. EPA. 2013) (see
Table 11-2 of this appendix). Since then, multiple new studies have added to this evidence base. Many
different factors are included in physiological stress, including oxidative stress, corticosterone (CORT)
levels, and immune response, all of which are discussed here.
Two different studies investigated CORT levels in response to Pb exposure. Meillere et al. (2016)
evaluated the relationship between feather Pb levels and feather CORT levels in wild common blackbirds
(Tardus merida) along an urbanization gradient. Male adult blackbirds were found to have an average
feather Pb concentration of 1.00 ± 0.76 |ig/g. dry weight, which was positively correlated with the degree
of urbanization. Feather CORT levels were found to be significantly and positively related to both the
degree of urbanization and feather Pb levels. Herring et al. (2018) also investigated CORT levels in birds.
Examining the relationship between fecal CORT levels (FCORT) and blood Pb levels in common ravens
(Corvus corax), it was found that blood Pb significantly affected FCORT levels only when there was
simultaneous exposure to mercury (Hg). FCORT was either not related or negatively correlated with
blood Pb when blood Hg concentrations were below 0.2 (ig/g, wet weight. Above this blood Hg
concentration, the FCORT response increased with increasing blood Pb concentrations.
Another aspect of physiological stress that has been linked to Pb exposure is oxidative stress.
Espin et al. (2014) assessed oxidative stress related to Pb in the Eurasian eagle owl (Bubo bubo). One
study in three different subareas in Murcia, southeastern Spain (rural, industrial, and mining areas)
evaluated the relationship between Pb exposure and oxidative stress biomarkers in blood. Glutathione
peroxidase (GPx) activity had a significant inverse correlation with Pb concentrations. Catalase (CAT)
activity was inversely related to Pb concentration as well. Both GPx and CAT are antioxidant enzymes
that catalyze the breakdown of free radicals and indirectly support the antioxidant defense system. Espin
et al. (2016) also examined these oxidative stress biomarkers in relation to blood Pb concentrations with
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different results. In two different gull species, Audouin's gull (Ichthyaetiis caidoiiinii) and slender-billed
gulls (Chroicocephcthis genei), total glutathione (GSH) content, antioxidant enzymes activities (GPx,
superoxide dismutase (SOD), CAT, GST), and lipid peroxidation (thiobarbituric acid reactive substances)
were analyzed to determine whether blood Pb concentrations had any effect on these oxidative stress
biomarkers. The only significant linear regression on Pb was the positive effect of Pb on GSH levels in
Audoin's gulls. The authors speculated that this could reflect the necessity to up-regulate GSH to balance
increased oxidative stress caused by metals. A laboratory study of female Japanese quail (Coturnix
japonica) also examined these effects, as well as other effects including liver histology and lipid
metabolism (Kou et al., 2020). Quail were fed one of five experimental concentrations of Pb solution (0,
50, 250, 500 and 1,000 ppm) for 49 days. Pb exposure of 250, 500, and 1,000 ppm induced severe
histopathological damages (liver lipid vacuoles and accumulation, hepatic cytoplasmic hyalinization and
vacuolization, hepatocyte necrosis, hepatic sinusoid congestion). It also led to a significant decrease in
GPx, SOD, and CAT activities in the liver.
Immune response has also been linked to Pb exposure, for example, in the following two studies.
Vermeulen et al. (2015) examined the effects of Pb exposure on the innate immunity of great tit (Parns
major) nestlings in populations along a metal pollution gradient. Average Pb concentration in red blood
cells was significantly higher in the populations closest to the pollution source than the farthest
population. There were significant differences in lysis scores among the populations, with lysis varying
inversely to Pb concentrations. Meissner et al. (2020) used the ratio of heterophils to lymphocytes (H/L
ratio) in mute swans (Cvgnns olor) to determine physiological stress levels. A higher H/L ratio indicates a
higher immune response, thus higher physiological stress. Mean blood Pb concentration was 0.239 (ig/g
(range: 0.028-0.675 (ig/g). H/L ratio was found to increase with blood Pb level, indicating that birds with
higher blood Pb levels had higher physiological stress.
11.2.4.4.2 Organism-Level Response
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a causal relationship
between Pb exposure and reproduction and developmental endpoints in terrestrial vertebrates (U.S. EPA,
2013) (see Table 11-2 of this appendix). Since the 2006 AQCD (U.S. EPA, 2006) and 2013 Pb ISA (U.S.
EPA, 2013), several field studies have examined the relationship between Pb exposure and reproduction.
Fritsch et al. (2019) found that the lifetime breeding success of free-living female European blackbirds
(Tardus menda) in Northwest Poland decreased with increasing levels of Pb in tail feathers
(average tail feather Pb = 6.7 |ig Pb/g dry weight). This same study also examined the relationship
between breeding success, lifespan, and Pb exposure. In birds with the greatest exposure and highest
breeding success, there is likely a trade-off between breeding effort and survival, as their lifespans tended
to decrease as Pb exposure increased. Chatelain et al. (2016) also studied how Pb exposure affected
reproduction. Adult feral pigeons (Columba livid) were dosed with one of four exposure treatments: Pb
only (1 ppm Pb acetate in tap water), Zn only (10 ppm ZnS04 in tap water), Pb and Zn (1 ppm Pb
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acetate + 10 ppm Zn sulfate in tap water), or control (tap water with no metal addition) every other day
for 2 weeks. One-day old nestlings of parents exposed to Pb (Pb and Pb +Zn groups) weighed
significantly less than the nestlings from other treatments (control and Zn groups) (mean 14.94 ± 0.72 and
17.20 ± 0.67 g, respectively). Additionally, eggs from parents exposed to Pb had significantly thinner
eggshells than those from the other groups (mean: 0.47 ± 0.00 and 0.49 ± 0.01 mm respectively).
While Fritsch et al. (2019) examined reproduction at the organism level, Hargitai et al. (2016)
examined suborganismal level responses to Pb exposure in relation to reproduction. Hargitai et al. (2016)
found that in great tit (Parns major) eggs from both woodland and urban habitats in the Pilis Mountains
of Hungary, egg yolk lutein and retinol levels were negatively related to the concentrations of Pb in the
eggshell. Lutein and retinol are both important antioxidants related to embryo viability in birds.
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a likely causal
relationship between Pb exposure and neurobehavioral effects in terrestrial vertebrates (U.S. EPA, 2013)
(see Table 11-2 of this appendix). Several additional studies in birds have since been published that assess
Pb effects on behavioral endpoints in birds. One new study of the neurobehavioral effects of Pb-exposure
evaluated the relationship between the behavior of free-living Northern mockingbirds (Mimns
polyglottos) and the soil Pb concentrations in their habitats in New Orleans, LA (McClelland et al., 2019).
Birds living in neighborhoods with high soil Pb concentrations had higher Pb concentrations in their
blood and feathers than those from the neighborhood with low soil Pb concentrations. This study used
simulated territory intrusions to examine the level of aggression displayed by individuals from different
neighborhoods. Birds from the high Pb neighborhoods exhibited a more aggressive response to simulated
intrusions than birds from the low Pb neighborhood.
Another study of the effects of Pb exposure on behavior examined how early-life dietary Pb
exposure in great tits (Parns major) affected both physical and neurological development (Ruuskanen et
al., 2015). Wild birds in selected nests were given an oral dose of Pb acetate in distilled water (4 (ig/g
body weight for high exposure and 1 |ig/g body weight for low exposure) every day for 12 days, starting
at 3 days after hatching. At 15 days old, the birds were brought into captivity and kept there for the
remainder of the experiment to assess their development after Pb exposure. Early-life Pb exposure was
found to have no effect on activity, exploration, neophobia, or success in learning and spatial memory
tasks.
Commonly observed effects of Pb on terrestrial vertebrates include decreased survival,
reproduction, and growth, as well as effects on development and behavior (U.S. EPA, 2006). The 2013
Pb ISA (U.S. EPA, 2013) also provided evidence for Pb effects on hormones and other biochemical
variables. New studies have expanded upon the relationship between Pb exposure and a-ALAD activity
by adding more species of birds, amphibians, and mammals to the evidence base. More evidence of
oxidative stress has been gathered, as well as evidence of effects on CORT levels and immunity in birds.
Literature since the 2013 Pb ISA continues to add to evidence relating to reproductive effects at both the
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organism and suborganism levels including effects on lifetime breeding success and some specific
secondary sexual traits. New studies of behavioral effects included increased aggression in mockingbirds.
11.2.5 Exposure and Response of Terrestrial Species
As previously reported in the (U.S. EPA, 1977), the 1986 Pb AQCD (U.S. EPA, 1986), the 2006
Pb AQCD (U.S. EPA, 2006) and the 2013 Pb ISA (U.S. EPA, 2013), a large number of experimental
studies have exposed a wide variety of terrestrial organisms to gradients of Pb exposures and reported a
broad assortment of responses, including growth, reproduction, survival, antioxidant levels and markers
of oxidative stress. More than 80 such additional experimental studies conducted since the 2013 Pb ISA
were identified. Organisms subjected to these exposure-response experiments have included various wild
plants including reeds and ferns, cultivated crops, microbes, lichens, fungi including mycorrhizae,
bacteria, nematodes, worms, collembolans, beetles, spiders, rodents, and birds. The 2006 AQCD and
2013 Pb ISA (U.S. EPA, 2013, 2006) reported that variation in exposure is generally associated with
commensurate variation in growth, reproduction, survival, antioxidant activity and more. Such coupling
of exposure and response is considered a strong indicator of causality (U.S. EPA, 2015), and exposure-
response studies with Pb thus continue to provide evidence supporting the causality of Pb for the effects
they investigate, as highlighted in the sections of this appendix dedicated to specific groups of terrestrial
organisms.
With very few exceptions, experimental exposure-response studies of terrestrial organisms
generate multiple level of exposure through addition of various soluble salts of Pb to the culture medium
(natural or artificial soil or hydroponic solution) or to food in the case of some animals. This makes it
possible to create a gradient that is easy to quantify and manipulate and is isolated from confounding,
nuisance and interacting variables. In principle, these attributes are desirable, as they allow for a more
accurate measurement and modeling of exposure-response relationships. They may introduce limits on
the scope of inference, but can nonetheless lead to credible, accurate predictive estimates of response,
within an acceptable range of natural conditions wherein factors other than exposure are left to vary
freely. However, in the particular case of terrestrial organisms and estimates of their response that are
obtained through experiments in which exposure is accomplished using salts of Pb, this may not be the
case. These experiments are informative for establishing causality, but not for deriving accurate predictive
estimates of response under natural conditions.
Section 11.2.2.1 discussed environmental variables that have a strong impact on bioavailability in
soils. They include pH, CEC, salinity, aging, OM, soil type and the presence of other metals. The use of
soluble salts of Pb brings pH, CEC, salinity, and aging into ranges far removed from those found in
natural environments following exposure to Pb emissions. Predicted effects derived from those
experiments cannot be expected to be accurate in environmental conditions, not only because the
experimental conditions of pH, CEC, salinity and aging diverge too far from those present in the
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environment, but, more intractably, because in both the experiments themselves and in the environments
in which a prediction is attempted, the measurement of Pb concentration may sharply diverge from the
concentration actually affecting the organism. These difficulties were discussed in the 2006 Pb AQCD
(U.S. EPA, 2006) and the 2013 Pb ISA (U.S. EPA, 2013), as well as in studies explicitly designed to
clarify these issues, such as Smolders et al. (2009), Cheyns et al. (2012) and Dayton et al. (2006). In
2009, following extensive toxicity testing with both spiked soils and contaminated field soils, Smolders et
al. (2009) concluded that "despite all of the efforts made, a large proportion of the difference between the
toxicity observed in field-contaminated soils and that in laboratory-amended soils remains unexplained."
Smolders et al. (2009), further demonstrated that not only are the effects of pH, for example, more
complex than previously thought, but pH, CEC, DOM, Fe and Mn oxides, aging and soil type are all
powerful modifiers of Pb toxicity to soil-dwelling organisms. Underscoring the complexity of modifying
effects, Cheyns et al. (2012), for instance, showed that in tomato and barley plants, soil type is a major
modifier of toxicity, but that once pH is controlled, toxicity can be mediated by the nutrient deficiencies
that stem from reactions of Pb with essential nutrients in the soil solution, whereby the apparent effects of
Pb are caused by nutrient deficiencies from Pb robbing the plants of P, for example, by forming Pb
phosphates.
The following more recent studies have continued to untangle the respective roles of the various
factors that complicate predictive estimation of the effects of Pb in terrestrial organisms from exposure-
response studies (Table 11-3). With enough knowledge of the effects of these factors on the exposure-
response relationship, it could, in principle, become possible to use some of those experiments to generate
useful estimates of concentrations associated with responses of interest from experiments. Experimental
procedures might be adjusted, for example by aging or leaching soils prior to exposing organisms, or the
modeling of the relationship itself might be modified, for example by adding correction coefficients to the
exposure.
Among other questions, Zhang and Van Gestel (2019b) investigated the effects of 18 months of
aging, form of Pb and percolation (leaching) on the toxicity of Pb to the worm Enchytraens crvpticiis in
natural standard soil spiked with nine levels of Pb between 0 and 3,200 mg Pb/kg dry soil using Pb(NC>3)2
and between 0 and 1,000 mg Pb/kg dry soil using PbO. Among the complex interactions between these
variables, they found that while leaching dramatically decreased porewater concentration of Pb in fresh
and aged soils, and more so for Pb(NC>3)2 than for PbO, it did not affect Pb uptake, which was greater for
the more soluble form (PtyNOsb). LC50 and LC10, estimated from logistic regression on all nine levels
was higher following leaching for Pb(N03)2 but not for PbO. The authors concluded that generally, the
effect of percolation on the toxicity of Pb-spiked soils was dependent on the chemical form used for
spiking as well as on aging, and porewater Pb concentration could not explain Pb toxicity. For survival,
leaching decreased the toxicity of Pb(N03)2 but did not affect the toxicity of PbO. For effects on
reproduction, leaching had a greater influence in freshly spiked soils than in aged soils. This suggests that
manipulating or accounting for aging and form of Pb might be useful in generating effect predictions in
natural environments from spiking experiments, but that manipulating leaching may not be.
11-77
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The same authors also included variation in the length of the aging period in another exposure-
response experiment (Zhang and Van Gestel. 2019a). Using the same materials and methods as in Zhang
and Van Gestel (2019b). they incubated the soil samples for five periods from 0 to 18 months, after
spiking and before exposure of the worms. Toxicity increased with aging when soils were spiked with
PbO but not with Pb(NC>3)2, as did availability when estimated via CaCh extraction. This may conflict
with (Smolders et al.. 2015). who found that lethality declined with five years of aging, but in outdoors
conditions that included leaching by rain rather than laboratory incubation. Including aging in the
translation from experiment to field thus appears warranted, but not without also including the form of Pb
and leaching.
Finally, Zhang et al. (2019a) investigated the effects of soil properties toxicity to Enchytraens
crypticus using the same materials and methods as Zhang and Van Gestel (2019b) and six standard
natural soils, using Pb(NC>3)2 but not PbO treatments, and not varying aging or percolation. The soils
varied in OM content, pH, CEC, water-holding capacity, dissolved OC, and composition. Soil type had
very large effects on survival of earthworms in the presence of Pb even though no effect was observed on
the internal Pb concentration of worms, with effects ranging from no survival at the midrange of Pb
concentration, to complete survival even at the highest concentration. However, soil type had only weak
effects on survival when exposure was measured as porewater Pb and no effect on survival when
measured as CaCh-cxtractable Pb. Similarly strong effects of soil type were seen on the exposure-
response relationship of Pb concentration and earthworm reproduction. However, the same weak effects
of Pb as for survival were observed for reproduction when using porewater Pb concentrations, and no
effects were observed when using CaCh-extractable Pb. Furthermore, measuring exposure as CaCh-
extractable Pb resulted in accurate and precise predictions of responses regardless of soil type. In
contradiction with other studies cited above, such as Chevns et al. (2012) or Smolders et al. (2009). the
authors suggested that despite soil type having a strong effect on toxicity when exposure is measured as
simple soil concentration using CaCh-extractable Pb as a metric of exposure may be sufficient when
estimating the effects of Pb on worms, since using that metric supported accurate and precise prediction
of earthworm responses regardless of soil type, and the exposure-response relationship was then
insensitive to soil type.
Romero-Freire et al. (2015) assessed the respective influence of soil properties in laboratory
toxicological assays, with the same aim of making experimental exposure-response studies with spiked
soils usable for environmental risk assessment. Seven natural soils of varying pH, conductivity, texture,
OC, water-holding capacity, CEC, specific area, carbonate content and metal oxides were spiked with five
levels ofPb(N03)2 and incubated for 4 weeks. The authors observed that pH and CaCO, content were the
soil properties with the highest influence on Pb extractability and interacted strongly with total Pb
concentration, with extractability most affected at higher concentrations of Pb. However, they also found
that retention via organic complexation kept most of the Pb from being bioavailable and that texture
(silt/sand/clay proportions) and Fe and Mn oxides also had major effects on extractability. In three tests of
toxicity—one with lettuce seeds, one with a strain of marine bacterium, and one measuring microbial soil
11-78
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respiration—soil type strongly modified overall toxicity in all tested organisms and the relative effects of
each concentration of Pb (in other words, the slope of the response curve). In addition, the magnitude of
these modifying effects differed among the three tests. The authors did not attempt to partition the effects
of every soil property beyond the most salient effects on extractability noted above. They concluded that
soil properties in the particular locations and land use where risk is to be assessed must be taken into
consideration when conducting risk assessment, including at minimum, pH, OM and carbonate.
Many variables distinguish natural soils from each other with regard to influence on Pb toxicity,
as enumerated in the experiments cited here. As noted by Romero-Freire et al. (2015). Zhang et al.
(2019a). Smolders et al. (2009) and others, given practical limitations on the number of soils that can be
included in one experiment, it is not possible to definitively separate the effect of each of the variables
that define soil type, let alone quantify their interactions. It is possible however to separate some variables
that affect the exposure-response more strongly from those that have little or no influence, and it may be
possible to identify measures of exposure under which the exposure-response relationship is insensitive to
soil type, but nonetheless support accurate and precise estimation of toxic effects.
Another study of the factors that contribute most strongly to differences between responses
occurring in natural environments and those observed in Pb-spiking experiments was conducted by
Smolders et al. (2015). The study was aimed at assessing the relative magnitude of the effects of salinity,
acidification, and aging on the toxicity of Pb to invertebrates, plants, and microbes. Samples of three
natural soils were spiked with seven levels of Pb ranging from 0 to 8,000 mg Pb/kg as Pb(NC>3)2 and as
PbCk Some samples were used unleached and imaged, some were leached and pH-corrected, and some
were leached, pH-corrected and aged for five years, a much longer period than in most aging studies.
Tomato and barley seedlings were grown in all nine treatments, and biomass was measured after 21 days.
Nitrification and soil respiration were measured to assess microbial activity, and the reproduction of the
worm E. fetida and the collembolan F. Candida was likewise measured for the nine treatments. Relative to
the unaged, unleached treatment, the increase in ECio with leaching and pH correction, aging or leaching,
pH correction and aging, showed very wide variation between endpoints. All endpoints demonstrated
strong toxicity relative to controls at all levels of added Pb in all three unaged, unleached soils. The EC50
for all endpoints increased with leaching and pH correction except for earthworm reproduction in one
soil, again with wide variation among endpoints. Finally, aging for five years combined with leaching and
pH correction increased EC50 to such a degree for all endpoints that its value could not be estimated for
any of them. Earthworm reproduction was the endpoint for which EC50 increased the least. Smolders et al.
(2015) attempted to identify which variables among total Pb concentration, porewater Pb, Pb2+ ionic
activity, pH and porewater ionic strength were most strongly correlated with endpoints. Overall,
porewater ionic strength was the variable most strongly correlated with toxicity. Based on this correlation,
the authors suggest that increased salinity, i.e., salt stress compounding true Pb toxicity in freshly spiked
soils, is likely the greatest modifying factor of toxicity. They found the effect of pH to be inconclusive
due to limitations of their experimental protocol, and perhaps surprisingly, caution about giving too much
weight to the effects of aging despite its seemingly large effect. They re-emphasized the limitations of the
11-79
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experimental protocol, specifically the leaching that preceded aging. For plants, they noted a deficiency of
P, with both increased Pb concentration and aging as the more direct factor explaining the effects on plant
growth. The authors concluded that regardless of the mechanisms behind their observations, this study
offered "... a strong confirmation that acute dosing of soluble Pb2+ salts does not appear to be an
appropriate model for environmental sources of Pb where Pb gradually enters soils via atmospheric
deposition as PbO, PbS, and PbSC>4..."
In 2021, Ports et al. (2021) proposed two corrections to the results of exposure-response
experiments conducted with addition of soluble salts of Pb to soil and used them to derive some examples
of ecological soil standards. They suggested first that a single correction factor can be applied to the
toxicity results of fresh, i.e., unleached, unaged, spiking experiments to adequately convert the results to
the values that would have been observed following leaching and aging. They further proposed to
demonstrate that this conversion generates values that correspond to the toxicity levels that would be
observed in corresponding hypothetical field conditions. The second correction was intended to adjust
differences in toxicity that arise from differing soil properties. Although as referenced previously,
multiple properties of soils have been shown to affect Pb toxicity in both spiking experiments and field
conditions, the authors argued that adjusting for CEC is sufficient. The authors demonstrated the
derivation of predicted no-effect concentrations (PNEC) according to the European REACH Regulation
European Parliament and Council (2006). using the two corrections above and data that conformed to the
REACH requirements. In contrast with Eco-SSL values, none of the derived standards were lower than
background soil Pb concentration.
A few methodological developments in analyzing and using Pb exposure-response experiments
have also been explored since the 2013 Pb ISA (U.S. EPA. 2013). although they may not be of immediate
applicability to risk assessment or standard setting. Zhang and Van Gestel (2017) used one standard
natural soil spiked with seven levels of Pb(NC>3)2 between 0 and 3,200 mg/kg soil to study the
toxicokinetics and toxicodynamics of uptake, elimination, and survival in the worm Encytraeus crypticus.
Uptake and toxicity were measured at seven time intervals and elimination at six. The measurement and
statistical modeling of the time course of uptake, elimination and survival demonstrated that accumulation
and toxicity were dependent on exposure duration, and that once the time course of exposure was taken
into consideration, the internal concentration of Pb in worms may be a better predictor of survival than
soil concentration. Using the model organism C. elegcms exposed to five levels of Pb between 0 and
2000 ppm as Pb acetate, Sudama et al. (2013) combined chromatographic metabolite profiling and
principal component analysis to show that changes in the purine pathway and its metabolites can be
detected after exposure to extremely low concentrations of Pb.
Finally, the applicability of Species Sensitivity Distribution analysis was investigated by Ding et
al. (2016) using 21 natural soils spiked with four levels of Pb between 0 and 350 mg/kg soil as Pb(NC>3)2
and 12 cultivars each of carrot (Daucus carota), radish (Raphanus sativus), and potato (Solatium
tuberosum), to show that Species Sensitivity Distribution analysis could be a reliable approach to
11-80
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determining safety thresholds, as long as the threshold values are derived from experiments designed for
that purpose. However, exposure was from soluble salt, and the safety thresholds the authors investigated
were for the safety of human consumers of vegetables grown in heavily polluted sites. They therefore
measured only accumulation in the plants and not the effects on the plants themselves.
11-81
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Table 11-3 Studies of factors that affect the interpretability of exposure-response experiments in terrestrial
biota, since the 2013 Pb ISA
Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional
study factors
Reference
Barley
(Hordeum
vulgare)
Tomato
(Lycopersicon
esculentum)
Form of
Pb:
PbCI2
jdH:
Soils 7.4, 6.5, 6.7, 5.7, 5.2, 4.7 (pH
CaCh (0.01 M) Adjusted with CaO)
Medium: Hydroponics 6.1
Six topsoils
from five
European
countries
CEC:
Soils 14.7, 27.1, 8.7, 4.2, 7.6, 41.7
(cmolc/kg soil)
Hydroponic
system
Exposure
method:
Salt mixed
with soil
Salt in
hydroponic
solution
Hydroponics N/A
PC:
Soils 14, 31, 10, 15, 21, 310 (gC/kg
soil)
Hydroponics N/A
Aaina/leachina:
Soils. Leached by immersion and
draining after 1 wk incubation. Three
1-wk periods of moist incubation
separated by 1-wk periods of dry
storage and one dry storage period
of up to 20 wk
Hydroponics N/A
Soils Measured:
6 levels * 6
soils = 36 values
between 47 and
12,700 mg/kg,
plus 1
control x 6 soils
with background
Pb between 4.7
and
135 mg Pb/kg
soil
Hydroponics 1,
3.2, 10, 32, 100,
320 mM
Soil
(location of
origin)
Soil P
content 44,
48, 67, 89,
90,
121 mg P/kg
soil
Hydroponics
Gradually
increasing P
supply for
17 days to
maintain
growth rate
and avoid
precipitation
AND
7 levels of P
supply
based on P
content in
plant tissue
(0.10%-
0.32% P in
plant tissue)
Tomato
growth:
decreasing
with
increasing
Pb in all soils
Barley
growth: no
effect of Pb
in three soils,
decreasing
with
increasing
Pb in three
soils
Tomato shoot
dry weight
NOEC for six
soils:
4,400, 750,
<250, 440, 260,
1,100 mg Pb/kg
soil
ECso: 6,000,
6,500, 2,200,
2,700, 1,600,
5,400 mg Pb/kg
soil
Barley shoot dry
weight NOEC for
six soils: >7,200,
>5,000, 2,000,
>3,400, 260,
1,100 mg Pb/kg
soil
ECso: >7,200,
>5,000, 4,900,
>3,400, 1,900,
8,300 mg Pb/kg
soil
Strong
interaction
effect of soil
type and Pb
concentration
on growth
P content in
plants was
strongly
influenced by
Pb and
explained the
effect of Pb
across soils and
in hydroponic
experiment
Chevns et
al. (2012)
11-82
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Study
Pb factors Effects of Effect Effects of
Organism Experimental conditions concentrations otheprbthan Pb concentration Reference
exposure
Potworm
Form of
(Enchytraeus Ekl
crypticus) Pb(N03)2
PbO
Medium:
LUFA2.2
standard
soil
Exposure
method:
Soil spiked
with
powdered
salt
Ehli
Nominal pH 5.49
CEC:
9.10 cmolc/kg
PC:
not reported
Aaina/leachina:
Spiked soils were aged for 0, 3, 6,
12 and 18 mo. No leaching
Nominal
concentrations
of:
Pb(N03)2
0, 50, 100, 200,
400, 600, 800,
1,600 and
3,200 mg Pb/kg
dry soil
PbO
0, 78, 156, 312,
625, 1,250,
2,500, 5,000 and
10,000 mg Pb/kg
dry soil
Aging and
chemical
form of Pb
E. crypticus
mortality
increased
with
increasing
Pb soil
concentration
Pb(N03)2:
CaCI2
extractable Pb
0, 3, 6, 12, 18-
mo LCso = 2.18,
3.06, 2.49, 2.28,
1.72 mg Pb/kg
ECso = 0.149,
0.125, 0.090,
0.103,
0.093 mg Pb/kg
Pore water Pb
0, 3, 6, 12, 18-
mo
LCso = 0.247,
0.346, 0.328,
0.366,
0.583 mg Pb/L
ECso = 0.020,
0.016, 0.019,
0.015,
0.046 mg Pb/L
The dose- (Zhang
response and Van
curves and Gestel.
toxicity values 2019a)
(LCso and EC50)
based on total
Pb
concentrations
differed widely
between the
two forms of Pb
Pb(N03)2 was
more toxic than
PbO in freshly
spiked soils, but
the toxicity of
PbO increased
with aging,
while the
toxicity of
Pb(NOs)2
remained
constant
Internal Pb
0, 3, 6, 12, 18-
mo LCso = 76.2,
76.4, 77.1, 73.4,
76.8 mg Pb/kg
dry body weight
ECso = 22.2,
24.7, 30.0, 31.5,
20.1 mg Pb/kg
dry body weight
CaCh-
extraction
provided the
best estimate of
Pb toxicity and
bioaccumulation
11-83
-------
Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional
study factors
Reference
PbO:
CaCI2
extractable Pb
0, 3, 6, 12, 18-
mo
LCso = 3.02,
3.15, 2.36, 2.66,
2.45 mg Pb/kg
ECso = 0.170,
0.135, 0.098,
0.138,
0.101 mg Pb/kg
Pore water Pb
0, 3, 6, 12, 18-
mo
LCso = 0.262,
0.312, 0.286,
0.302,
0.391 mg Pb/L
ECso = 0.025,
0.048, 0.050,
0.023,
0.048 mg Pb/L
Internal Pb
0, 3, 6, 12, 18-
mo LCso = 78.0,
77.7, 74.4, 78.4,
78.7 mg Pb/kg
dry body weight
11-84
-------
Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional
study factors
Reference
ECso = 23.7,
18.1, 19.8, 16.9,
12.0 mg Pb/kg
dry body weight
Potworm Form of
(.Enchytraeus Ekl
crypticus) Pb(NC>3)2
PbO
Medium:
jdH:
Aged Soil:
p H pw. 5.61
pHcace: 5.14
LUFA2.2 Freshy Spiked:
standard
soil
pHpw: 5.93
pHcace: 5.65
CEC:
powdered nQt ^
salt.
Exposure
method:
Soil spiked
with
PC:
not reported
Aqinq/leachinq:
One soil form was spiked and then
aged for 18 mo, while the other soil
form was used without aging as
freshly spiked soil
Half of each set of soils were
leached with deionized water equal
Nominal
concentrations
of:
Pb(N03)
0, 50, 100, 200,
400, 600, 800,
1,600 and
3,200 mg Pb/kg
dry soil
PbO
0, 78, 156, 312,
625, 1,250,
2,500, 5,000 and
1,000 mg Pb/kg
dry soil
Percolation,
chemical
form of Pb
and aging
E. Crypticus
mortality
increased
with
increasing
Pb soil
concentration
PbfNOsh:
CaCI2-
extractable Pb
aged,
aged+leached,
freshly spiked
and freshly
spiked+leached.
LCso = 1.72,
2.42, 2.07 and
2.78 mg Pb/kg
ECso = 0.093,
0.173, 0.044 and
0.109 mg Pb/kg
Pore water Pb
aged,
aged+leached,
freshly spiked
and freshly
spiked+leached.
LCso = 0.583,
0.201, 0.686 and
0.148 mg Pb/L
ECso = 0.046,
0.063, 0.012 and
0.033 mg Pb/L
When exposure (Zhang
was measured and Van
as total soil Pb, Gestel.
aging increased 2019b)
toxicity for both
forms of Pb and
leaching had no
meaningful
effect. However,
all effects of
form, aging or
leaching
disappeared
when exposure
was measured
as CaCb-
extractable Pb
11-85
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Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional Reference
study factors
to two times the base moisture
content
Internal Pb
aged,
aged+leached,
freshly spiked
and freshly
spiked+leached.
LCso = 76.8,
84.4, 77.3 and
83.6 mg Pb/kg
dry body weight
ECso = 20.1,
22.1, 25.5 and
32.7 mg Pb/kg
dry body weight
PbO:
CaCI2
extractable Pb
aged,
aged+leached,
freshly spiked
and freshly
spiked+leached.
LCso = 2.45,
2.01, 2.79 and
2.16 mg Pb/kg
ECso = 0.101,
0.160, 0.123 and
0.168 mg Pb/kg
Pore water Pb
aged,
aged+leached,
freshly spiked
11-86
-------
Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional
study factors
Reference
and freshly
spiked+leached.
LCso = 0.391,
0.233, 0.197 and
0.097 mg Pb/L
ECso = 0.048,
0.043, 0.047 and
0.031 mg Pb/L
Internal Pb
aged,
aged+leached,
freshly spiked
and freshly
spiked+leached.
LCso = 78.7,
76.4, 83.3 and
84.5 mg Pb/kg
dry body weight
ECso = 12.0,
14.5, 41.1 and
38.6 mg Pb/kg
dry body weight
Potworm
Form of
jdH:
Nominal
Exposure
Toxicity was
Days 4, 7, 10, 14
Strong
(Zhana
(Enchytraeus
Pb:
Nominal pH of 5.49
concentrations
duration
dependent
and 21
interaction
and Van
crypticus)
Pb(N03)2
of 0, 100, 200,
on both the
effect of
Gestel.
400, 800, 1,600
concentration
Total
duration and Pb
2017)
CEC:
and
and duration
concentration
Medium:
9.10 cmolc/kg
3,200 mg Pb/kg
of exposure.
concentration in
soil:
on mortality
LUFA2.2
dry soil
standard
soil
OC:
Pb toxicity
LCso = 2,336,
2278, 1,220, 756
not reported
Measured
Concentrations
of 16, 114, 202,
developed
more slowly
than uptake,
and
558 mg Pb/kg
dry soil
11-87
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Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional
study factors
Reference
Exposure
method:
Soil spiked
with
aqueous
salt solution
Aaina/leachina:
After 14-d exposure in spiked soils,
the surviving E. crypticus were
transferred to clean soil for the 14-d
elimination phase
391, 793, 1,601
and
3,585 mg Pb/kg
dry soil
with final
LCso not yet
reached after
21 d
Internal
concentration:
LCso = >287,
>270, 161, 76.6
and
76.4 mg Pb/kg
dry body weight
Potworm
criticus)
Form of
(Enchytraeus
Pb(N03)2
Medium:
Five
standard
soils (LUFA
standard
soil 2.1, 2.2,
2.3, 2.4,
5 M) and
one soil
from a
soccer field
in the
Netherlands
Exposure
method:
Soils spiked
with
aqueous
solution
jdH:
4.86, 5.66, 5.38, 6.87, 6.99, 6.85
CEC:
2.23, 7.59, 4.04, 20.1, 10.1 and
20.0 cmolc/kg
PC:
DOC
45.7, 61.7, 34.4, 72.0, 51.2 and
189 mg/L
Aging/leaching:
Soil equilibrated for 14-d.
Nominal
concentrations
of 0, 100, 200,
400, 600, 800,
1,200, 1,600,
2,400 and
3,200 mg Pb/kg
dry soil
Soil type,
soil
properties:
OM, DOC,
PH, CEC,
water-
holding
capacity,
composition
Reproductive
toxicity and
mortality
increased
with Pb
concentration
in soil
LUFA standard
soil 2.1, 2.2, 2.3,
2.4, 5 M and
soccer field,
respectively.
Total Pb:
LCso = 246,
1,192, 655,
3,125, 2,875 and
>3,092 mg Pb/kg
dry soil
ECso = 81.4,
238, 205, 948,
1,008 and
991 mg Pb/kg
dry soil
CaCI?
extractable Pb:
LCso = 2.35,
2.11, 1.86, 1.64,
2.11 and
>1.39 mg Pb/kg
dry soil
Correlation of
single soil
properties with
endpoints,
followed by
simple
regression,
followed by
stepwise
multiple
regression
suggested that
pHcaci2 was the
best
explanatory
factor for LCso
values based
on total Pb
concentration
The differences
between soil
toxicity were not
present when
exposure was
measured as
CaCh-
extractable Pb
concentration
(Zhang et
al„ 2019a)
11-88
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Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional Reference
study factors
ECso = 0.329,
0.193, 0.107,
0.180, 0.241 and
0.115 mg Pb/kg
dry soil
Porewater Pb:
LCso = 0.308,
1.25, 0.335,
0.334, 0.933 and
>0.754 mg Pb/L
ECso = 0.044,
0.127, 0.117,
0.169, 0.046 and
0.105 mg Pb/L
Internal Pb:
LCso = 95.7,
83.0, 87.0, 84.3,
81.7 and
>47.7 mg Pb/kg
dry body weight
ECso = 13.6,
34.1, 26.0, 39.9,
27.2 and
32.6 mg Pb/kg
dry body weight
Tomato
Form of
eUi
Nominal
Leaching
All effects
ECsos calculated
Strong
Smolders
(Lycopersicon
Pb:
6.1-7.4
concentrations
combined
increased
for tomato
interaction
et al.
esculentum)
PbCI2
of, 250, 500,
with pH
with
growth, barley
effect of
(2015)
CEC:
1,000, 2,000,
correction,
increasing
growth,
leaching, aging
4,000 and
aging
Pb in freshly
nitrification rate,
and Pb
Pb(N03)2
8.2-27.1 cmolc/kg soil
8,000 mg Pb/kg
combined
spiked
nitrification 28-d,
concentration
with
(unaged,
respiration, E.
on all
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Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional
study factors
Reference
Barley
(Hordeum
vulgare)
Collembola
(Folsomia
Candida)
Earthworm
(Eisenia
fetida)
Medium:
Soils
gathered
from
topsoils in
Spain, the
United
Kingdom
and Belgium
Exposure
method:
Soil spiked
with salt
PC:
10-43 g C/kg soil
Aqinq/leachinq:
Soils were given three different
treatments. Treatment A: freshly
spiked. Treatment B: leached and
pH-corrected. Treatment C: leached,
pH-corrected and aged for 5 yr
leaching
and pH
correction
unleached)
soils
Fetida
reproduction and
F. Candida
reproduction in
each soil,
respectively.
Spain:
Freshly spiked
ECso = 2,900,
2,380, 3,240,
7,190, 8,720,
480 and
712 mg Pb/kg
soil
Leached and
pH-corrected
ECso = 6,370,
7,190, 2,200,
7,120, 12,300,
1,182 and n.s.
mg Pb/kg soil
responses.
Leaching
combined with
pH correction
decreased
toxicity for all
effects. Aging
following
leaching and pH
correction
further
decreased
toxicity for most
effects but not
all. Authors
suggest
decreased ionic
strength (salt
stress) and
changes in pH
are the main
drivers of
decreasing
toxicity
Aged 5 yr
ECso = 12,600,
n.s, n.s, n.s,
7,020, 1,270 and
n.s. mg Pb/kg
soil
United Kingdom:
Freshly spiked
ECso = 6,140,
6,750, 2,820,
1,750, 9,970,
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Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional
study factors
Reference
2,400 and
4,530 mg Pb/kg
soil
Leached and
pH-corrected
ECso = 6,420,
5,020, 4,920,
n.s., 6,160,
1,700 and
5,020 mg Pb/kg
soil
Aged 5 yr
ECso = n.s., n.s.,
n.s., n.s., n.s.,
3280 and n.s.
mg Pb/kg soil
Belgium:
(no test for E.
fetida)
Freshly spiked
ECso = 1,240,
1,710, 1,470,
1,410, 1,680 and
1,710 mg Pb/kg
soil
Leached and
pH-corrected
ECso = 1,430,
4,580, 1,640,
2,820, 8,150 and
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Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional
study factors
Reference
2,700 mg Pb/kg
soil
Aged 5 yr
ECso = 4480,
n.s., n.s., n.s.,
n.s. and n.s.
mg Pb/kg soil
Lettuce
(Lactuca
sativa)
Bacterium
(Vibrio
fischeri)
Form of
Pb:
Pb(N03)2
Medium:
Seven soils
representing
the main
soil groups
in Spain
Exposure
method:
Spiked with
aqueous
solution.
Reported for soils H1-H7
respectively
Ehli
7.96, 8.67, 8.79, 6.74, 7.20, 5.87
and 7.03
CEC:
21.4, 9.83, 2.94, 9.91, 25.9, 3.83
and 15.5 cmolc/kg
PC:
5.43, 0.42, 0.38, 0.61, 8.22, 0.49
and 0.66%
Aqinq/leachinq:
Soils were incubated for 4 wk after
spiking
Nominal
concentrations
of 500, 1,000,
2,000, 4,000 and
8,000 mg Pb/kg
soil
Soil type
(location of
origin)
All effects
increased
with
increasing
Pb in all soils
Reported for
soils H1-H7,
respectively
L. sativa'.
EC10 = 499,
1,363, 254,
1,097, 3,452,
498 and
344 mg Pb/kg
soil
V. fischeri:
EC10 = >8,000,
5,337, 2,901,
386, 2,473, 8
and
744 mg Pb/kg
Soil Respiration:
EC10 = >8,000,
3,128, 5,951, 90,
>8,000, 122 and
45 mg Pb/kg
Strong
interaction
effect of soil
type and Pb
concentration
on all
responses.
Authors suggest
that the main
soil properties
that affected
toxicity were
pH, carbonate
content and OC
Romero-
Freire et
al. (2015)
CaCI2 = calcium chloride; CaO = calcium oxide; CEC = cation exchange capacity; DOC = dissolved organic carbon; EC50 = half maximal effect concentration; LC50 = 50% lethal
concentration; LUFA = Landwirtschaftliche Untersuchungs- und Forschungsanstalt; mo = months; N/A = not available; NOEC = no-observed-effect concentration; n.s. = nonsignificant.
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Organism
Experimental conditions
Pb
concentrations
Study
factors
other than
Pb
exposure
Effects of
Pb
Effect
concentration
Effects of
additional
study factors
Reference
OC = organic carbon; OM = organic matter; P = lead; Pb = lead; Pb(N03)2 = lead nitrate; PbCI2 = lead chloride; PbO = lead(ll) oxide; pHpw = pH of porewater; pHCaci2 = pH via calcium
chloride; wk = week(s); yr = year(s).
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11.2.6
Terrestrial-Community and Ecosystem Effects
In the 2013 Pb ISA the body of evidence was sufficient to conclude there is a likely causal
relationship between Pb exposure and terrestrial-community and ecosystem effects (U.S. EPA, 2013). In
the 2006 Pb AQCD (U.S. EPA, 2006), terrestrial ecosystems near stationary Pb sources exhibited
decreased species diversity, changes in floral and faunal composition, and a reduction in vegetation
fitness. In the 2013 Pb ISA (U.S. EPA, 2013), a study reported decreased population growth of
earthworms. Additional studies in the 2013 Pb ISA examined how the presence of AMF or earthworms
affect plant Pb uptake and fitness. Recent evidence of the effects of Pb at the community and ecosystem
levels includes several studies of the relationship between Pb soil concentration and species interactions
and invertebrate community structure. Specifically, studies conducted since the 2013 Pb ISA have
reported that Pb affects plant-insect interactions and is correlated with invertebrate community structure.
Considering that Pb rarely occurs as the only contaminant in terrestrial ecosystems it is difficult to
attribute effects observed at higher levels of biological organization solely to Pb.
In an experimental study, Jiang et al. (2020) demonstrated trophic transfer of Pb can affect the
chemical defenses of larch seedlings (Larix olgensis) against an economically important pest, the Asian
gypsy moth (Lvmantria dispctr), in China. Larch seedlings were enriched with Pb at 0, 500, or
1500 mg Pb/kg. Second instar L. dispctr larvae raised from field-collected egg masses were placed on L.
olgensis seedlings for 7 days. Pb content in L. dispctr larvae were significantly higher than L. olgensis
needles for the 500 mg Pb/kg and 1500 mg Pb/kg treatments, and Pb bioaccumulated in this experiment,
as the transfer coefficients were 0.97 for the 0 mg Pb/kg treatment, 5.43 for the 500 mg Pb/kg treatment
and 6.03 for the 1500 mg Pb/kg treatment. Pb treatment reduced L. olgensis total biomass (40.36%
reduction in the 1500 mg Pb/kg compared with control) and L. dispar larval weights (by 34.44%-
52.05%) and survival rates (by 30.91%-59.28%) in a dose-dependent manner compared with the control.
Antioxidants (peroxidase and SOD) of L. olgensis increased under 500 mg Pb/kg treatment and were
reduced under 1500 mg Pb/kg. Phytochemical defenses, protease inhibitors (trypsin inhibitor and
chymotrypsin inhibitor) and the secondary metabolites (total phenolic acids) were significantly increased
under the low dose of Pb (500 mg Pb/kg) compared with the control, while all phytochemical defense
chemicals, including condensed tannins, decreased significantly under high Pb stress (1500 mg Pb/kg).
Lvmantria. dispctr fed with L. olgensis seedlings had higher antioxidase activities in the fourth instar
(SOD and CAT), while nonenzymatic antioxidants were significantly decreased (glutathione content and
ascorbic acid content), suggesting that the reduction of antioxidants might lead to the oxidative stress
experienced by L. dispar larvae. Finally, MDA content increased with Pb exposure.
Heavy-metal concentration along a pollution gradient in Romania affected soil mite (Acari:
Mesostigmata) community structure (Manu et al., 2019; Manu et al., 2017). Manu et al. (2017) examined
soil mite communities in relation to soil metal content and physicochemical properties in 12 grasslands.
Some heavy metals (Pb, As, Cu and Zn) influenced the soil mite community in highly polluted sites,
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while altitude and soil humidity played larger roles in less polluted sites. Pb soil concentration ranged
from 28.21 ± 4.62 mg/kb Pb to 421.12 ± 71.62 mg/kb Pb. The sites with the highest Pb were closest to the
pollution source. Canonical correspondence analysis (CCA) determined that heavy metals (Cu, Zn and
Pb) as well as the C/N ratio, humidity, total N, altitude, and slope were the strongest determinants of
species composition, and Pb soil concentration showed association with the abundance of Zercon
berlesei. In another study, Manu et al. (2019) collected soil from a pollution gradient surrounding the
Certej ore deposit and characterized heavy-metal concentration and soil mite communities. Pb
concentrations ranged from 153.68 to 292.35 mg Pb/kg across five sites (mean concentration). The
relationship between mite abundance and heavy metals was examined using CCA, and the first axis
accounted for 50.67% of the variation in mite community and was highly correlated with Pb
(correlation = 0.81), Cu, As and Mn. The abundance of Arctoseius cetratus showed the strongest
relationship with Pb.
Potworm (Enchytraeidae) diversity, but not herbaceous plant diversity, was negatively correlated
with soil Pb concentration across 41 sites near a Zn-Pb mining site in South Poland (Kapusta and
Sobczvk. 2015). Pb soil concentration varied across sites, ranging from 300 ± 300 mg Pb/kg
(mean ± S.D.) to 9,600 ± 14,100 mg Pb/kg at sites closer to the smelter, and water-soluble Pb showed a
similar pattern, with higher Pb concentrations found closer to the smelter site (range:
0.103 ± 0.068 mg Pb/kg to 0.477 ± 0.212 mg Pb/kg Pb). Pb concentration was positively correlated with
silt content, OC, total Cd, total Zn, exchangeable Cd, water-soluble Cd, and water-soluble Zn and
negatively correlated with distance from the smelter. Water-soluble Pb was positively correlated with
distance from the smelter, OC, and water-soluble Zn. Total Pb was significantly negatively correlated
with Enchytraeid species richness, genus richness, and density in 2010, but not density in 2009, while
water-soluble Pb showed no significant relationships with species richness, genus richness, or density in
2009 or 2010. Plant community species richness and herbaceous cover showed no correlation with total
Pb in the soil or water-soluble Pb.
The abundance of insects on Pb-contaminated kale (Brassicct oleracea L. var. acephala) was
higher than control B. oleracea plants in a field experiment in Brazil (Morales-Silva et al.. 2022).
Brassica oleracea plants were grown in control soil (background Pb concentration: 25.9 mg Pb/kg) or in
soil spiked with Pb(NC>3)2 to nominal concentrations of 144, 360, or 600 mg Pb/kg and exposed to natural
insect populations. Lepidoptera and their associated parasitoids, as well as aphids and their predators and
parasitoids, were collected from plants. At the end of the experiment, plant biomass was unaffected by Pb
soil contamination, while plants exposed to 600 mg Pb/kg had significantly higher concentrations of Pb in
the leaves compared with plants in the control, 144, and 360 mg Pb/kg treatments. Brassica oleracea
plants in the control treatment had significantly higher abundance of insects compared with the
contaminated plants, regardless of Pb level.
Longer-lived nematodes with lower fecundity are most affected by experimental Pb exposure
(Park et al.. 2016). Tomatoes (Lycopersicon escidentiim) were grown in pots of soil collected from an
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agricultural field in Korea and exposed to Pb via irrigation. Measured Pb concentrations of the soil were
16.97 ± 0.24 mg Pb/kg (mean ± S.D) for the control soil, 15.19 ± 0.55 mg Pb/kg, 15.54 ± 0.42 mg Pb/kg,
18.08 ± 0.67 mg Pb/kg and 34.98 ± 2.57 mg Pb/kg. Soil nematode communities were characterized before
L. esculentum were planted and after 18 weeks of growth. Nematode community structure was analyzed
using a variety of metrics, from trophic guilds to maturity indices to the abundance of colonizers and
persister (cp-1 = colonizer to cp-5 = persister). Pearson's correlation coefficients between Pb and
nematode community indices were largely nonsignificant, except for the negative relationship between Pb
and the richness of cp-3 as well as the maturity index and the positive relationship between Pb and the
abundance of fungivores as well as the abundance of cp-2. There was a significant decrease in nematode
abundance in omnivores-predators (OP) and cp-4 at the highest concentrations of Pb. Nematode richness
decreased at higher concentrations of Pb, particularly for OP, cp-4, and cp-5. The authors suggested that
these groups are likely most sensitive to environmental stress, as they have longer-lifecycles and lower
reproduction rates.
In another nematode study, the diversity and abundance of nematode communities were
correlated with soil Pb concentration near a ferroalloy manufacturer in North Slovakia (Salamun ct al..
2011). Soil samples near the factory and downwind of the factory were analyzed for heavy metals,
including Pb. The total Pb concentration ranged from 0.815 ± 0.471 mg Pb/kg to 1.766 ± 0.082 mg Pb/kg
(mean ± S.D). Soil Pb concentration was positively correlated with the abundance of certain trophic
guilds and ecological indices of nematodes, specifically, predators, root-fungal feeders, and maturity
index (MI2-5). Maturity index (2-5) is used as a measure of functional diversity, which incorporates the
abundance of r and /^-strategists in a community. Pb was not significantly correlated with any other
trophic group or ecological index (bacterial feeders, fungal feeders, omnivores, plant feeders, maturity
index, plant-parasite index, genera richness, Shannon-Weaver index, Simpson index or abundance). In a
follow-up study, Salamun et al. (2012) examined nematode community structure in relation to the total
element concentration of Pb, Zn, Cu Cr, Ca, and As, in another region of Slovakia using an HNO3
extraction and mobilization fraction Na2EDTA extraction. Unlike Salamun et al. (2011). in which Pb was
positively correlated with certain trophic groups, total soil Pb concentration was negatively correlated
with the abundance of omnivorous nematodes, MI2-5, structure index and genera richness.
Since the 2013 Pb ISA (U.S. EPA. 2013). several studies have found evidence that Pb affects
species interactions, including chemical defenses (Jiang et al.. 2020) and pollinator foraging behavior
(Xun et al.. 2018). Additionally, several studies found negative relationships between Pb concentration
along a pollution gradient and aspects of the invertebrate community structure, specifically in soil mites
(Manu et al.. 2019; Manu et al.. 2017). potworms (Kapustaand Sobczvk. 2015). insect communities on
kale (Morales-Silva et al.. 2022). and nematodes (Salamun et al.. 2011).
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11.3
Freshwater Ecosystems
11.3.1 Summary of New Information on Effects of Pb in Freshwater
Ecosystems and Causality Determination Update Since the 2013 Pb ISA
Recent evidence further supports the findings of the previous Pb AQCDs and 2013 Pb ISA that
waterborne Pb is toxic to freshwater plants, invertebrates, and vertebrates, with toxicity varying with
species and lifestage, duration of exposure, form of Pb, and water quality characteristics (U.S. EPA,
2013, 2006, 1986, 1977). In natural environments it is difficult to attribute observed effects solely to Pb
due to the presence of confounding factors such as other pollutants, and additional modifying factors that
affect Pb bioavailability and toxicity. Furthermore, the portion of Pb from atmospheric sources is usually
not known. The majority of the available studies of Pb exposures in freshwater biota are laboratory
toxicity tests on single species in which an organism is exposed to a known concentration of Pb, and the
effect on a specific endpoint is evaluated. These studies provide evidence for a temporal sequence
between Pb exposure and an effect, an aspect important in judging causality. Concentration-response data
from freshwater organisms indicate that there is a gradient of response to increasing Pb concentration
and that some effects in sensitive species are observed at or near the upper limit of Pb concentrations
quantified in U.S. surface waters (Table 11-1). New evidence for freshwater biota (Table 11-5) continue
to support the existing causality determinations from the 2013 Pb ISA summarized in Table 11-4 of this
document. In most cases, new evidence expands somewhat the evidence for endpoints that were already
established as causal in the 2013 Pb ISA. Some studies have reported effects at lower effect
concentration than in the 2013 Pb ISA. There are no changes to existing causality determinations for
freshwater biota or ecosystems from the 2013 Pb ISA (Table 11-4).
For physiological stress endpoints in freshwater plants, invertebrates, and vertebrates, new
evidence continues to support the likely to be causal determination from the 2013 Pb ISA. A small subset
of studies that report molecular or cellular perturbations of Pb concurrently assess an effect on
reproduction, growth, or survival. Few studies were identified since the 2013 Pb ISA that quantified
ALAD response in freshwater invertebrates or vertebrates; hence there is not sufficient evidence to
warrant a reconsideration of any of the causality relationships for the hematological effects of Pb.
Neurobehavioral effects of Pb were concluded to have a likely to be causal relationship for Pb
exposure for freshwater invertebrates and vertebrates in the 2013 Pb ISA. For invertebrates, a few new
studies in amphipods, bivalves and gastropods further support the 2013 finding of a likely to be causal
relationship between Pb exposure and neurobehavioral endpoints (Section 11.3.3). Effects on locomotion
were observed in adult amphipods, G. fossarum, following Pb sublethal exposure (analytically verified
concentrations were 2.1 and 2.7 |ig Pb/L in two separate studies, one conducted for 24 hours, another
conducted for 5 days) (Lebrun and Gismondi, 2020; Lebrun et al„ 2017). Alteration of neurotransmitter
(AChE) activity was reported for two freshwater bivalve species including Parreysia corrugcita, in which
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AChE activity was significantly induced at 26 |ig Pb/L in 21-day aqueous exposure. Impaired foot
movement was also observed in this species at a similar concentration (Brahma and Gupta. 2020). AChE
activity was significantly induced in the freshwater snail B. aeruginosa during 28-day exposure to Pb-
spiked sediment (29.7 mg Pb/kg dry weight) (Liu et al.. 2019b).
The 2013 conclusion of a likely to be causal relationship between Pb exposure and
neurobehavioral effects in freshwater vertebrates is bolstered in this current ISA by multiple studies with
zebrafish (D. rerio) as an animal model for human health effects including developmental and
neurological changes associated with Pb exposure (Section 11.3.4.4). Effects on behavioral endpoints
such as locomotion and social interactions in larval zebrafish were reported at lower effect concentrations
than studies in the 2013 Pb ISA, with some effects reported at < 20 |ig Pb/L; a subset of these studies
analytically verified Pb in the exposure water (Kataba et al.. 2020; Zhao et al.. 2020; Wang et al.. 2018b;
Zhu et al.. 2016). Neurological responses of fish to Pb exposure were first reported in the 1986 Pb AQCD
(U.S. EPA. 1986). The likely to be causal determination in the 2013 Pb ISA was based primarily on
altered behaviors, such as reduced locomotion and prey capture ability, observed in fish following Pb
exposure. These included a decrease in zebrafish larval startle response to mechanosensory and visual
stimuli following nominal exposure to Pb (2.0 and 6.0 |ig Pb/L) (Rice et al.. 2011). and reduced prey
capture in assays with 10-day old fathead minnows born from adult fish exposed to 120 |ig Pb/L for
300 days then subsequently tested in a breeding assay for 21 days (Mager et al.. 2010). In another study in
the 2013 Pb ISA with fathead minnows, swimming performance measured as critical aerobic swim speed
was significantly impaired in minnows in 24-hour acute (139 (ig Pb/L) and chronic 33 - to 57-day
(143 |ig Pb/L) exposures; however, no significant difference in swim speed was observed in chronic
exposures to 33 |ig Pb/L (Mager and Grose 11. 2011). The evidence in the 2013 Pb ISA and previous
AQCDs also included effects on molecular targets; however, these experiments were typically conducted
at Pb concentrations that greatly exceeds environmental concentrations.
In the 2013 Pb ISA, there was a conclusion of a causal relationship between Pb exposure and
reduced survival in both freshwater invertebrates and vertebrates. Newly available evidence continues to
support these causal determinations. For invertebrates, several studies provide further characterization for
known effects on survival in a few sensitive species of freshwater invertebrates at <20 |ig Pb/L
(Section 11.3.5). In the gastropod L. stagnalis, survival was significantly decreased at 8.4 |ig Pb/L after
21-day exposure to the end of a 56-day full lifecycle assessment (Munlev et al.. 2013). In a chronic 42-
day bioassay with the amphipod H. azteca, the EC20 for survival was similar under two different
experimental diets administered concurrently (LC20 =15 jag Pb/L and LC20 =13 jag Pb/L) (Besser et al..
2016). For freshwater vertebrates, studies in fish provided the basis for causality determination in the
2013 Pb ISA (Section 11.3.5). Additional fish bioassays conducted in varying water chemistry conditions
report effects on survival at Pb concentrations similar to those reported in the 2013 Pb ISA. For larval
zebrafish (D. rerio), 96-hour LC50 values varied with water hardness; LC50 = 52.9 |ig Pb/L in soft water
and LC50 = >590 |ig Pb/L in hard water (Alsop and Wood. 2011). Several studies considered the role of
Pb and other trace metals on the decline of the white sturgeon in U.S. waters, and one study examined
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endpoints in westslope cutthroat trout. In 96-hour acute toxicity assays conducted with two lifestages of
white sturgeon (A. trcmsmontamis), the lowest 96-hour LC50 was 177 |ig Pb/L for 8 dph larvae (Vardv et
al.. 2014).
For growth effects in freshwater organisms associated with Pb exposure, recent studies continue
to support the findings in the 2013 Pb ISA. There was a likely to be causal relationship between Pb
exposure and reduced plant growth concluded in the 2013 Pb ISA. Most primary producers experience
EC50 values for growth at concentrations that greatly exceed Pb concentrations typically found in U.S.
surface waters. One new study reported growth rates in three commonly tested algal species (P.
subcapitata, C. kesslerii, and C. reinhardtii) at lower effect concentrations than previously reported. P.
subcapitcita was the most sensitive in 72-hour bioassays, with an EC50 = 83.9 |ig Pb/L,
EC20 = 45.7 |ig Pb/L and EC10 = 32.0 |ig Pb/L based on filtered Pb concentration. Varying the pH resulted
in greater sensitivity (Dc Schamphelaere et al.. 2014). In the 2013 Pb ISA, there was a causal relationship
concluded to exist between Pb exposure and reduced growth in invertebrates. Since then, additional
studies have supported previous findings of Pb effects on the growth of snails (L. stagnctlis) in the
low (ig Pb/L range (Cremazy, 2018, 6708984} (Munlev et al.. 2013; Brix et al.. 2012; Esbaugh et al..
2012). Reduction in weight gain and specific growth rate were observed in juvenile Oriental river prawn
(M nipponense) exposed to 25 |ig Pb/L in chronic 60-day trials. No growth effects were observed in
prawns at 12 (ig Pb/L (Ding et al.. 2019). The evidence remains inadequate to infer a causality
relationship for Pb exposure and reduced growth in freshwater vertebrates. One study reported a threshold
of 160 |ig Pb/L for tadpole growth in dark-spotted frogs (P. nigromaculata) (Huang et al.. 2014).
Reproductive and developmental effects were concluded to be causally related to Pb exposure for
freshwater invertebrates in the 2013 Pb ISA. This remains the case in newer studies. Recent evidence
further supports previous observations of Pb effects on reproductive endpoints at low |ig/L concentrations
in sensitive species of gastropods, cladocerans and rotifers, especially under chronic exposure scenarios
(Section 11.3.5) (see Table 11-5). InZ. stagnctlis, a gastropod known to be sensitive to Pb at low |ig Pb/L
concentration, NOEC <1.0 (ig Pb/L and LOEC = 1.0 |ig Pb/L were determined for the number of egg
masses and time until the first egg mass in a 56-day lifecycle bioassay (Munlev et al.. 2013). In this
species, the egg capsule and embryo diameter were significantly reduced after 7 days of development at
2.7 (ig Pb/L (the highest concentration in which reproduction was observed in the study). For the
cladoceran C. dubia, 7-day EC20 values for reproduction ranged from 12 to 223 |ig Pb/L in assays
conducted in a variety of natural waters across the United States with different water chemistries; 7-day-
EC50 values ranged from 20 to 573 |ig Pb/L in the same test waters (Esbaugh et al.. 2012). Using the same
sampled waters from across the United States, reproduction (as population growth) was also assessed in
rotifer P. rapida over a 4-day exposure period. Chronic EC20 and EC50 in this species ranged from 3 to
103 |ig Pb/L and from 10 to 154 (ig Pb/L, respectively.
Several studies in fish in which Pb concentration was analytically verified further support the
causal determination reported in the 2013 Pb ISA between Pb and reproductive and developmental effects
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for freshwater vertebrates (Section 11.3.4.4). For example, hatching success rates in zebrafish embryos
were reduced at 4.5, 9.6 and 18.6 |ig Pb/L aqueous exposure; at 72 hpf, the hatching success rates at all
three concentrations were significantly decreased compared with the control, indicating that Pb caused a
hatching delay. This effect persisted until the end of the experiment at 96 hpf (Zhao et al., 2020).
Endocrine disruption (significant reduction in thyroid hormones T3 and T4) was observed in zebrafish
larvae following exposure to 30 (ig Pb/L, although there was no effect on the hatching success rate (Zhu
et al.. 2014). These studies in fish are bolstered by several analytically verified studies in amphibians
(Section 11.3.4.4.3).
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a likely to be causal
relationship between Pb exposure and freshwater-community and ecosystem effects, and recent evidence
continues to support this finding (Section 11.3.6). Reductions in species abundance, richness or diversity
associated with the presence of Pb in freshwater habitats are reported in the literature, usually in heavily
contaminated sites where Pb (and other metal) concentrations are higher than typically observed
environmental concentrations. Most evidence is from sediment-associated macroinvertebrate
communities. Observational and experimental studies published since the 2013 Pb ISA continue to show
negative associations between sediment and/or porewater Pb concentration and macroinvertebrate
communities. The evidence is expanded somewhat with studies reporting associations with Pb and
periphyton abundance. Uptake of Pb into aquatic and terrestrial organisms and subsequent effects on
mortality, growth, developmental and reproduction at the organism level can cascade up to ecological
populations and communities and lead to ecosystem-level consequences, and thus provide consistency
and plausibility for causality in ecosystem-level effects. Although the evidence is strong for the effects of
Pb on growth, reproduction, and survival in certain species in experimental settings at or near the range of
Pb concentrations reported in surveys of U.S. freshwater systems, considerable uncertainties exist in
generalizing effects observed at a smaller scale, particular conditions up to predicted effects at the
ecosystem level of biological organization. In many cases, it is difficult to characterize the nature and
magnitude of effects and to quantify relationships between ambient freshwater concentrations of Pb and
ecosystem response due to the presence of multiple stressors, variability in field conditions and
differences in Pb bioavailability at that level of organization.
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Table 11-4 Summary of Pb causality determinations for freshwater plants,
invertebrates, and vertebrates
Level Effect Freshwater3
2013 Pb ISAb
2024 PbISA
Community and Ecosystem
Community and Ecosystem Effects
Likely Causal
Likely Causal
Reproductive and Developmental
Effects - Plants
Inadequate
Inadequate
Reproductive and Developmental
Effects - Invertebrates
Causal
Causal
Population-
level
Reproductive and Developmental
Effects -Vertebrates
Causal
Causal
Growth - Plants
Likely Causal
Likely Causal
Endpoints
Organism-level
Growth - Invertebrates
Causal
Causal
Responses
Growth - Vertebrates
Inadequate
Inadequate
Survival - Plants
Inadequate
Inadequate
Survival - Invertebrates
Causal
Causal
Survival - Vertebrates
Causal
Causal
Neurobehavioral Effects -
Invertebrates
Likely Causal
Likely Causal
Neurobehavioral Effects - Vertebrates
Likely Causal
Likely Causal
Hematological Effects - Invertebrates
Likely Causal
Likely Causal
Hematological Effects - Vertebrates
Causal
Causal
Suborganismal
Responses
Physiological Stress - Plants
Likely Causal
Likely Causal
Physiological Stress - Invertebrates
Likely Causal
Likely Causal
Physiological Stress - Vertebrates
Likely Causal
Likely Causal
Conclusions were based on the weight of evidence framework for causal determination in Table II of the ISA Preamble (U.S. EPA.
2015)1. Ecological effects observed at or near Pb concentrations measured in sediment and water in Table 6-2 of the 2013 Pb ISA
were emphasized and studies generally within one to two orders of magnitude above the reported range of these values were
considered in the body of evidence for freshwater (Section 6.4.12) (U.S. EPA. 2013).
Inputs of Pb into freshwater ecosystems include air-related sources and non-air sources
(Appendix 1: https://assessments.epa.gov/isa/document/&deid=359536). Atmospherically derived Pb can
enter aquatic systems through direct wet or dry deposition and erosional transport or resuspension of Pb
from terrestrial systems (Section 11.1.2). Receiving water bodies include lakes (lentic systems) and rivers
and streams (lotic systems). Freshwater wetlands, some of which may be inundated occasionally or
constantly, also provide habitat for aquatic biota. The focus of this section is on Pb bioavailability,
bioaccumulation, and the effects of Pb on freshwater organisms including algae, aquatic plants, microbes,
invertebrates, vertebrates, and other biota with an aquatic lifestage (e.g., amphibians).
The following sections review the recent literature published since the 2013 Pb ISA on effects of
Pb on freshwater ecosystems. The new evidence is considered along with the ecological findings of
previous Pb assessments. The 2013 Pb ISA developed causality determinations for freshwater biota based
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on the weight of evidence for Pb effects on specific endpoints and taxonomic groups (Table 11-4). In the
2013 Pb ISA, the body of evidence was sufficient to conclude that there was a causal relationship between
Pb exposure and reproductive and developmental effects in freshwater invertebrates and vertebrates,
reduced growth and survival of invertebrates, reduced survival of vertebrates, and hematological effects
in vertebrates. Relevant concentrations for causality judgments for the welfare effects of Pb in the 2013
Pb ISA were determined considering Pb concentrations "generally within one or two orders of magnitude
above those which have been observed in the environment and the available evidence for concentrations
at which effects were observed in plants, invertebrates, and vertebrates" (U.S. EPA. 2013). Of these
causal relationships concluded for freshwater ecosystems, effects on reproduction, growth, and survival in
sensitive freshwater invertebrates are well characterized from controlled studies at concentrations at or
near Pb concentrations occasionally encountered in U.S. fresh surface waters. The 2013 Pb ISA
concluded there is a likely to be causal relationship between Pb exposure and physiological stress in
freshwater biota. For hematological effects, there was a likely to be causal relationship for freshwater
invertebrates. Effects on neurobehavioral endpoints were likely to be causal for freshwater invertebrates
and vertebrates. Pb effects on plant growth were likely to be causal and were only reported at relatively
high concentrations compared with effects on invertebrates. There was also a likely to be causal
relationship between Pb exposure and community and ecosystem-level effects. For all effects in
freshwater biota, the toxicity of Pb varied with species and lifestage, duration of exposure, form of Pb,
and water quality characteristics. Key uncertainties from the last review for freshwater ecosystems
included the uncertainties associated with generalization of effects observed in controlled laboratory
studies to conditions in streams, rivers, and lakes where many modifying factors affect Pb bioavailability
and toxicity. For example, there is a discrepancy between the sensitivity of aquatic insect taxa in
laboratory studies compared with longer-term field studies. In a meta-analysis of study findings, longer-
term studies suggest that aquatic insect taxa are more sensitive to metals than indicated in acute exposure
scenarios (Brix et al.. 2011). In aquatic ecosystems affected by Pb, exposures are most likely
characterized as low-dose, chronic exposures, whereas the majority of available toxicological data for this
metal is from acute laboratory exposures, typically conducted at higher concentrations. There are
considerable uncertainties associated with generalizing effects observed in controlled studies to effects at
higher levels of biological organization. Furthermore, available studies on community and ecosystem-
level effects are usually from contaminated areas where Pb concentrations are much higher than typically
encountered in the environment and multiple contaminants are present. At the time of the 2013 Pb ISA,
the connection between air concentration of Pb and ecosystem exposure was poorly characterized for
aquatic habitats (U.S. EPA. 2013). Furthermore, the previous review noted that the level at which Pb
elicits a specific effect is difficult to establish in freshwater systems, due to the influence of other
environmental variables (e.g., pH, OM) on both Pb bioavailability and toxicity, and due to substantial
species differences in Pb sensitivity. Evidence indicated that Pb is bioaccumulated in biota; however, the
sources of Pb in freshwater organisms have only been identified in a few studies, and the relative
contribution of Pb from all sources, including atmospheric deposition, is usually not known.
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Studies published since the 2013 Pb ISA that characterize bioavailability and uptake of Pb, and its
effects in freshwater organisms and ecosystems, that identify additional uncertainties, or decrease
uncertainties identified in the prior NAAQS review of this criteria air pollutant are presented throughout
the following sections. Brief summaries of conclusions from the 1977 Pb AQCD (U.S. EPA, 1977), the
1986 Pb AQCD (U.S. EPA. 1986). the 2006 Pb AQCD (U.S. EPA. 2006) and the 2013 Pb ISA (U.S.
EPA, 2013) are included where appropriate. Recent research on the bioavailability and uptake of Pb into
freshwater organisms including plants, invertebrates and vertebrates is presented in 11.3.2. Information on
environmental concentrations in freshwater biota and ecosystems in the United States at different
locations and overtime is presented in Section 11.3.3. Toxicity of Pb to freshwater flora and fauna
including growth, reproductive and developmental effects (Section 11.3.4) are followed with data on the
exposure and response of freshwater organisms (Section 11.3.5). Responses at the community and
ecosystem levels of biological organization are reviewed in Section 11.3.6.
11.3.2 Factors Affecting Bioavailability, Uptake and Bioaccumulation and
Toxicity in Freshwater Biota
Toxicity of Pb to aquatic life varies with the physicochemical properties of surface waters (U.S.
EPA, 2013). Factors affecting the bioavailability and subsequent toxicity of Pb to biota include chemical
factors (primarily water hardness, DOC, pH) and biological factors (e.g., lifestage, development of
tolerance, organism interactions). Water hardness, DOC, and pH can be quantified, are directly related to
the toxic effects and are used in bioavailability models to predict acute and chronic toxicity (Adams ct al..
2020) (Section 11.1.6). Biological factors discussed in prior Pb AQCDs or the 2013 Pb ISA that may
influence organism response to Pb exposure include the lifestage of an organism, genetics, and nutrition
(see Section 7.2.3, 2006 AQCD (U.S. EPA, 2006) and Section 6.4.9, 2013 Pb ISA (U.S. EPA. 2013)).
These factors are more difficult to link quantitatively to toxicity. Often, species differences in
metabolism, sequestration and elimination rates control the relative sensitivity and vulnerability of
exposed organisms. The organism route of exposure also influences Pb toxicity. Uptake of Pb by aquatic
invertebrates and vertebrates may preferentially occur via exposure routes other than direct absorption
from the water column, such as ingestion of contaminated food and water, uptake from sediment
porewater, or incidental ingestion of sediment (U.S. EPA, 2013, 2006). Fewer studies assess uptake,
bioaccumulation, and subsequent toxicity of Pb via diet than via aqueous exposure. Of the available Pb
feeding studies in freshwater biota, only a few pair the same concentration of waterborne exposure with
dietary exposure to compare the relative importance of dietary versus aqueous uptake pathways (Alsop et
al., 2016; DeForest and Meyer, 2015). Studies published since the 2013 Pb ISA on chemical factors
(water hardness, DOC, pH, temperature, and other metals) and biological factors discussed in this section
further enhance understanding of Pb uptake and subsequent toxicity in freshwater systems. Biological
factors include those that were well characterized in previous AQCDs and the 2013 Pb ISA,
(e.g., lifestage), and factors not previously considered, such as the role of parasites in modulating Pb
bioaccumulation.
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11.3.2.1.1 Water Hardness
The role of water hardness (the amount of Ca2+ and Mg2+ ions) in Pb uptake and subsequent
toxicity was reported in previous Pb AQCDs and the 2013 Pb ISA. Furthermore, U.S. EPA's existing Pb
AWQC are hardness-based (Section 11.1.7.3) (U.S. EPA. 1985a). Generally, as water hardness increases,
there is less Pb uptake due to competition of Ca2+ and Mg2+ for binding sites. Newer literature has
continued to examine the role of Ca2+ and Mg2+ and other cations commonly present in surface waters
(e.g., K+, Na+) in modulating Pb bioaccumulation and toxicity. For example, in a study of the amphipod
Gammcirus pulex exposed 2 days to 10 (ig Pb/L and a range of environmentally relevant cation
concentrations (Na+, Mg2+ or Ca2+), both Na+ and Mg2+ had no significant effect on Pb uptake while
increasing Ca2+ concentrations inhibited Pb uptake (Urien et al.. 2015). In a study reviewed in the 2013
Pb ISA, Ca2+ influenced Pb accumulation and toxicity in the fathead minnow (Pimephales promelas)
during waterborne exposure (Grosell et al.. 2006a). In a newer study in fish, Ca2+, Mg2+ or H+
significantly decreased Pb accumulation and toxicity in zebrafish larvae Danio rerio, while K+ and Na+
showed no effect (Feng et al.. 2018) (see Section 11.3.2 for further discussion of water hardness and Pb
toxicity).
As described in prior AQCDs and the 2013 Pb ISA, the effect of water hardness is variable;
generally, both the acute and chronic toxicity of Pb increase with decreasing water hardness as Pb
becomes more soluble and bioavailable and less Ca2+ and Mg2+ ions are available to compete with Pb for
binding sites. Studies available since the 2013 Pb ISA are also illustrative of the varying influence of
water hardness on the toxicity of Pb. In reproductive toxicity tests with C. dubia, 7-day EC50 was
81.2 (ig Pb/L at 10 mg/L Ca (0.25 mM) and 130 |ig Pb/L at 70 mg/L Ca (1.75 mM), showing that the
daphnids tested in the soft water were more sensitive to Pb toxicity (Nvs et al.. 2014). However, in a
bioassay with the rotifer Brctchionus calyciflorus, Ca was not protective in a chronic (48-hour) exposure
(Nvs et al.. 2016b). In bioassays with zebrafish larvae, Pb was more toxic in soft water
(11.7 mg CaCCh/L) compared with hard water (141 mg CaCCh/L) (Alsop and Wood. 2011).
11.3.2.1.2 Dissolved Organic Matter and Dissolved Organic Carbon
In studies cited in the 2013 Pb ISA, DOC was shown to have a protective effect on Pb toxicity in
freshwater invertebrates and fish Esbaugh et al. (2011); Mager et al. (201 la); Mager et al. (201 lb), and
newer studies continue to support these observations. Esbaugh et al. (2012) compared the relative
importance of water chemistry variables including DOC, Ca, and pH in the toxic response of freshwater
cladoceran (Ceriodaphnia dubia), mollusk (Lvmnaea stagncdis) and rotifer (Philodina rctpida) to a range
of Pb concentrations in bioassays conducted in a variety of natural waters from across North America.
The greatest toxicity to the cladoceran and snail species was observed in low-DOC waters, and toxicity
was found to be correlated with DOC using multilinear regression modeling analysis. This was not the
case in rotifer/', rapida, where toxicity was most closely correlated with Ca and pH, not DOC. In
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contrast, in the rotifer B. calycifloras, high DOC was protective against Pb chronic reproductive toxicity;
however, when expressed as free-ion activity, toxicity increased with increasing fulvic acid concentration
(Nvs et al.. 2016b). The authors suggest that fulvic acid-Pb complexes may also contribute to Pb
bioavailability in B. ccdycifloriis. Taking metal speciation into consideration, Dong et al. (2014)
calculated the Comparative Toxicity Potential of Pb (described as the ecotoxicological impact associated
with a unit emission of substance to defined ecological receptors via different pathways of exposure). Pb
had the highest Comparative Toxicity Potential in water with low DOC, moderate pH and hardness, and
the lowest Comparative Toxicity Potential in water with moderate DOC, high pH, and hardness. Pb
typically has high affinity to DOC, resulting in a low fate factor (residence time) and bioavailability factor
(fraction of truly dissolved metal within total metal) (Dong et al.. 2014). Additionally, Zhang et al. (2021)
found that modeling Pb and other heavy metals was improved when incorporating total OC and AVS.
Since the 2013 Pb ISA, studies have further elucidated the relationship between the
characteristics of humic substances and Pb bioavailability, such as molecular weight (MW) or other
additional effects associated with solar irradiation. In lake sediments, Pb-humic acid complexes are more
stable when the MW of the humic acid is lower. In particular, humic acids with MW lower than 10 kDa
could increase the biosorption capacity of Pb (Bai et al.. 2019). While Pb-humic acid complexes are
discussed in the 2013 Pb ISA, the study by Kostic et al. (2013) suggests a mechanism for the binding of
Pb to humic acid may be the "acid-like" nature of Pb(II). Pb(II)-ions strong affinity for humic acid may be
explained by its borderline acid properties and by how humic acids behave as weak acid polyelectrolytes.
Humic acids carry a variety of oxygen-containing functional groups such as carboxylic, hydroxyl,
phenolic and carbonyl groups with oxygen as a donor atom, which helps them form strong bonds with
Pb(II). This is also supported by the study by Liu et al. (2022). which found Pb(II) caused greater
quenching (the decrease of fluorescence by the metal addition) in humic-like DOM compared with
protein-like DOM. The finding was likely due to humic-like components complexing with Pb(II) through
carboxyl and hydroxyl (-COOH and -OH) groups, which generally bonds to Pb(II) preferentially over
protein-like DOM that contains significant amounts of the amino group (-NH2).
The bioaccumulation capacity for Pb in algae is influenced by the presence of organic acids. Que
et al. (2020) found that adding organic acids, such as malic acid or citric acid prolonged the adsorption
equilibrium time of the algae-Pb binary system. Citric acid showed a greater bioaccumulation capacity for
Pb in algae than malic acid, due to ternary complex formation. The binding capacity of Pb to OM is also
influenced by solar or UV-B radiation. Pb complexation with representative humic substances (Suwannee
River humic acid and Suwannee River fulvic acid) decreased with increasing simulated solar radiation
(Spierings et al.. 2011). This may be due to an increase in the relative abundance of the carboxyl groups
in the photoaltered humic substances and from decreased aromaticity (and thus less electronegativity)
with increasing irradiation doses. The presence of Pb2+ can also increase the photodegradation of
microcystin and thus reduce microcystin accumulation in sediments and in certain fish (Dai et al.. 2017).
Reduced amounts of humic acid were adsorped to the freshwater microalga Chlorellct kesslerii, which
then reduced Pb bioavailability to the microalgae because the humic substances increase the
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bioavailability of Pb to microalgae by adding supplementary binding sites and because Pb uptake by C.
kesslerii is controlled by transport across the biological membrane rather than by diffusion in the medium
(Spierings et al.. 2011). However, there was no correlation with an increase in free Pb ions and algal
intracellular Pb content, likely due to the formation of additional binding sites on the photoaltered humic
acids. In additional tests using Elliott humic acid under simulated solar radiation, free Pb ions were
released from the metal-DOM complex as the irradiation dose increased, and there was a 33% increase in
intracellular Pb concentration in Chlamydomoncts reinhardtii at high irradiance (Worms et al.. 2015).
11.3.2.1.3 pH
As described in prior AQCDs and the 2013 Pb ISA, uptake and subsequent toxicity of Pb to
freshwater biota can be affected by pH, either directly or indirectly. Generally, at low pH, there is more
Pb2+ available to bind to the biotic ligand. As pH increases, there is increased formation of Pb organic
(DOC) and inorganic (OH-, COr ) complexes, which decrease Pb bioavailability. Since the 2013 Pb
ISA, several studies have further characterized Pb complexation and adsorption under changing pH
conditions. There are more binding sites for Pb to humic acids at pH 6 than at pH 4, likely due to the
higher content of dissociated functional groups in humic acids at higher pH, and more favorable
electrostatic attraction when binding surfaces become deprotonated at higher pH (Bai et al.. 2019). Xu et
al. (2018) found that the binding dynamics of DOM groups in response to Pb(II) addition were regulated
by both pH and ionic strength. Specifically, at lower pH and ionic strength (e.g., pH 4.7 and ionic strength
0.01 M), as Pb(II) was added, aryl C-H and carboxyl C = O groups gave the fastest response, followed by
polysaccharide C-OH and chromophoric groups at 265 nm (CDOM265). However, when pH was raised
to 6.0, the opposite binding sequence was found, in that the CDOM265 group was bound first, followed
by the polysaccharide C-OH and carboxyl C = O, and finally the aryl C-H groups. Hua et al. (2013) found
that Pb absorption to biofilms was greatest at pH 9, which was 3.5 times greater than that at the minimum
adsorption (pH = 7).
Several studies since the 2013 Pb ISA have tested the effects of changing pH on Pb toxicity to
biota. In the freshwater algal species Pseiidokirchneriella subcapitata, as pH increased from 6.0 to 7.6,
the 72-hour EC50 decreased from 72.0 to 20.5 |ig filtered Pb/L (De Schamphelaere et al.. 2014). Further,
Antunes and Kreager (2014) observed greater toxicity (more bioavailability) for common duckweed (L.
minor) at higher pH; this was due to less H+ and competition at the macrophyte binding sites. The
apparent increase in Pb2+ toxicity at pH >7.0 coinciding with a changing ratio of [Pb2+]/[Pb(OH)+] (due to
the marked increase in [Pb(OH)+]) suggests that Pb(OH)+ also contributed to the toxicological response.
In some freshwater invertebrates, recent studies generally support previous understanding that
higher pH is protective; however, these findings vary by the duration of the toxicity bioassays and by
taxa. In a series of chronic reproductive toxicity tests with daphnia C. dubia conducted at different pH
values, high pH was protective of Pb toxicity. At the lowest pH tested (pH 6.4), the EC50 = 99.8 |ig Pb/L,
while at the highest pH (pH 8.2), the EC50 = 320 |ig Pb/L (Nvs et al.. 2014). Similarly, decreasing toxicity
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of Pb to D. magna with higher pH was observed by Oin et al. (2014); as pH increased from 5.0 to 9.0, the
24 h-LCso increased from 784 |ig Pb/L to 9,473 |ig Pb/L, and the predicted proportion of free Pb2+ ion was
99.75% at pH 5.0 and 2.9% at pH 9.0. High pH was also protective in chronic reproductive toxicity tests
with rotifer B. calycifloriis. Both the population growth rate and population size generally decreased with
increasing pH in bioassays conducted at pH values ranging from 6.4 to 8.2 (Nvs et al.. 2016b). Wang et
al. (2016b) found that for crustaceans, Pb toxicity increased with increasing pH, but for mollusks and
worms, toxicity decreased with increasing pH. For fish, toxicity was least at neutral pH and increased at
lower or higher pH levels. The toxicity of Pb can increase at higher pH when there is less competition
between H+ and metal binding sites on cell-surface ligands. However, there may be higher toxicity at
lower pH due to increased solubility and altered Pb speciation, which can increase Pb bioavailability for
certain animals. Uptake studies in natural environments have also pointed to the importance of pH in
uptake of Pb. A field study conducted in 36 headwater streams in the Lake District of England reported
statistically significant correlations between total dissolved Pb in stream water and body burdens in the
sampled aquatic insect taxa (Leuctra spp., Simuliidae, Rhithrogena spp., Perlodidae) (De Jonge et al..
2014). In the streams, H+ ion activity was the overriding factor influencing Pb body burden, while DOC
was not a significant factor.
In fish, the effects of pH on toxicity were variable in studies cited in the 2013 Pb ISA. For
example, lower pH was shown to result in increased sensitivity to Pb in juvenile fathead minnows
following 30-day exposure to Pb at varying concentrations (Grosell et al.. 2006a). Additionally, Birceanu
et al. (2008) determined that fish (specifically rainbow trout) were more susceptible to Pb toxicity in
acidic, soft waters, characteristic of sensitive regions in Canada and Scandinavia. Hence, fish species
endemic to such systems may be more at risk from Pb contamination than fish species in other habitats. In
a study published after the 2013 Pb ISA, Esbaugh et al. (2013) compared three methods used to acidify
laboratory bioassay water on LC50 values in fathead minnow. Pb toxicity varied significantly depending
upon the acidification method used in the experiment. The authors recommended direct acid-base addition
rather than CO2 or 3-(N-morpholino)propanesulfonic acid buffer. In an approach that linked metal
accumulation with toxicity through a BLM-aided toxicokinetic-toxicodynamic model, Gao et al. (2015)
demonstrated that increasing concentrations of H+ in test media significantly reduced Pb accumulation in
zebrafish larvae within the exposure duration of >4-72 hours. In the same study, increasing [H+]
significantly decreased the mortality of the larvae at >12-96 hours.
11.3.2.1.4 Water Temperature
In the 2013 Pb ISA, water temperature was noted as a factor affecting the toxicity of Pb to aquatic
organisms, with higher temperatures generally leading to greater response; a few recent studies reported
variable responses to Pb with temperature. Isopods Asellus aquations exposed for 10 days to one of two
water temperatures (15 ± 1°C and 20 ± 1°C) and three concentrations of Pb (0.0353 (imol/L, 7.3 |ig Pb/L),
0.353 |imol/L (73 |ig Pb/L) and 0.882 (imol/L (181 |ig Pb/L) exhibited distinct responses at the two
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temperature treatments (Van Ginneken et al., 2019). At 15°C, respiration decreased as Pb concentration in
the isopods increased. In the higher temperature treatment, feeding and respiration rates were higher and
were positively correlated with Pb uptake and accumulation. Park et al. (2020) assessed survival,
malformation and heart rate in zebrafish embryos exposed to three analytically verified concentrations of
Pb (2, 10 and 17 |ig Pb/L) at two temperatures (26°C and 34°C). At 26°C, the survival rate decreased
early in the 7-day exposure at the two highest concentrations, reaching 73% at 10 |ig Pb/L and 57% at
17 |ng Pb/L by the end of the experiment, with no significant effect at 2 |ig Pb/L. At 34°C, the survival
rate decreased significantly in all concentrations and to a greater extent in the highest concentration; at
7 days, embryo survival at 17 (ig Pb/L was 30% that of the control. Malformations such as spinal
curvature were observed in all tested concentrations at both temperatures. At 34°C, heart rate was
significantly decreased at all Pb concentrations, while at 26°C, heart rate was significantly decreased at
the two highest tested concentrations.
11.3.2.1.5 Other Metals
Multiple metals are present simultaneously in aquatic environments and may interact with one
another, influencing Pb uptake and resulting in antagonistic, synergistic, or other toxic effects. Recent
advances in in multimetal research since the 2013 Pb ISA have included development and evaluation of
bioavailability models to predict the toxicity of acute and chronic metal mixtures, of which Pb is one
component (Nys et al„ 2017; Farley et al„ 2015; Santore and Ryan, 2015). Since the 2013 Pb ISA,
considerable research beyond the scope of this document (Section 11.1.1) has focused on metal mixture
assessment, including how uptake and bioaccumulation are affected in freshwater biota in the presence of
multiple metals. The mechanisms of metal interactions may include competition for the same metal
transporter at the biological membrane or displacement of one metal by another metal on DOM, which
leads to changes in the free metal ion concentration in water (Cremazy et al„ 2019). The effects of metals
on Pb uptake and toxicity vary by metal. In the juvenile freshwater snail L. stagnctlis, Ni and Zn had no
effect on Pb uptake, but a small but significant inhibitory effect was observed with Ag (Cremazy et al„
2019). In the isopod A. aquations exposed to Cd and Pb simultaneously, synergistic interactions occurred
with metal uptake as well as on growth rates and mortality rates when compared with single-metal studies
(Van Ginneken et al„ 2015). In juvenile rainbow trout (Oncorhvnchus mvkiss) uptake studies of binary
mixtures with Pb paired with other metals, Pb uptake into gill tissue was significantly inhibited in a
noncompetitive manner by Ag, Cd and Cu, while Ni and Zn had no effect on Pb uptake (Brix et al.,
2017). In another study with juvenile rainbow trout, there was no effect on ionoregulation at a low Pb
concentration of 5.4 |ig Pb/L (26.1 nmol/L) (Clemow and Wilkie, 2015). However, in combination with
Cd, there was greater-than-additive toxicity, likely due to differences in the underlying mechanism of
action, with some shared binding sites between the two metals. In 5-day postfertilization zebrafish larvae
exposed to Pb alone (10 (ig Pb/L), Cd alone (5 |ig Pb/L) or Pb + Cd since 4 hours postfertilization, the
respective mean concentrations of Pb and Cd in tissue were statistically significantly lower in the co-
exposure group than in the groups exposed to Pb or Cd alone (Liao et al., 2021). There were differences
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in behavioral outcomes in the three treatment groups; Pb primarily affected locomotor activity, Cd
affected circadian behavioral rhythm and the two compounds in combination were antagonistic for both
locomotor activity and behavioral rhythm. The bioavailability of Pb is also affected by the formation of
complexes with various Fe (oxyhydr)oxides, such as ferrihydrite, schwertmannite, jarosite, goethite,
hematite, and magnetite (Shi et al.. 2021). Fe (oxyhydr)oxides influence the speciation, partitioning and
transport of Pb through adsorption and coprecipitation, and this can vary by acidity, alkalinity,
temperature, and oxic conditions.
11.3.2.1.6 Lifestage
The differential sensitivity of early lifestages of aquatic biota to contaminants is well-established
in the scientific literature, such that national and international entities (e.g., U.S. EPA, Organisation for
Economic Co-operation and Development, European Union) have standardized laboratory toxicity assay
protocols that call for testing with embryo, larval or juvenile organisms to assess effects at the most
sensitive lifestages. Differences in susceptibility to Pb at distinct lifestages for freshwater invertebrates
and fish are discussed in Section 6.4.9.4 of the 2013 Pb ISA. Recent studies conducted with freshwater
organisms reviewed in Sections 11.3.4 and 11.3.5 continue to demonstrate that lifestage is an important
determinant of increased sensitivity to Pb. For example, endangered white sturgeon (Acipenser
transmontanns) were three and a half times more sensitive when exposed to Pb at 8 days posthatch (dph)
than at 40 dph (Vardv et al.. 2014).
11.3.2.1.7 Species Sensitivity
As described in previous U.S. EPA reviews of Pb, sensitivity to this metal can vary by several
orders of magnitude across freshwater biota. Pb elicits responses in some species at low (<5 to 10 |ig Pb/L
range under some water conditions) concentrations while others appear to be unaffected at concentrations
greatly exceeding 1,000 |ig Pb/L. In a study reported in the 2013 Pb ISA, a series of SSD showed the
greatest sensitivity to Pb in crustaceans, followed by cold water fish, and warm water fish and aquatic
insects, which exhibited a similar sensitivity (Brix et al.. 2005). A comparison of cladoceran and copepod
freshwater species curves generated by Wong et al. (2009) indicated that cladoceran species, as a group,
were more sensitive to the toxic effects of Pb than were copepods, with respective hazardous
concentration values for 5% of the species of 35 and 77 |ig Pb/L. Following the 2013 Pb ISA, Deforest et
al. (2017) used acute and chronic toxicity data across a range of freshwater species and genera, taking into
account the differences in sensitivity to Pb, to propose updated aquatic life AWQC for Pb
(Section 11.3.5).
Some uncertainty is associated with the extrapolation of toxicity values generated from
laboratory-based single-metal acute exposure assays to chronic exposure to multiple metals and other
contaminants in field studies. (Brix et al.. 2011) provided examples of acute laboratory exposures with
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aquatic insects that suggested the insects are relatively insensitive to metals, in contrast to field studies
that report sensitivity. The authors conducted a meta-analysis of laboratory and field studies that generally
supported the finding of greater sensitivity of aquatic insects in chronic exposure field conditions.
However, the majority of available field studies involve multimetal exposures. The authors speculated
there could be a difference in the mechanism of toxicity between acute exposure and chronic exposure in
aquatic insects or that dietary metal exposure is another important contributing factor to toxicity in these
organisms.
11.3.2.1.8 Development of Tolerance
Tolerance to prolonged Pb exposure may develop over time in some organisms as they
physiologically adapt and survive under low variations of various environmental stresses, including Pb.
Evidence for genetic selection in the natural environment has been observed in some aquatic populations
exposed to metals in studies, as reviewed in the 2006 AQCD. Fewer laboratory-based assays have
examined the development of Pb tolerance. In a study reviewed in the 2013 Pb ISA, multigenerational
exposure to Pb appears to confer some degree of metal tolerance to Chironomns plumosus larvae;
however, metal-tolerant larvae were significantly smaller than larvae reared under clean conditions
(Ye dam an ik am and Shazilli. 2008). In a more recent multigenerational test with D. magna exposed to an
analytically verified concentration of 50 |ig Pb/L, the LC50 ( = 430 |ig Pb/L at the F0 generation)
increased to 2,110 (ig Pb/L in the F9 generation. The LC50 of control organisms in the F9 generation
varied from 430 |ig Pb/L to 890 |ig Pb/L suggesting that the Pb-exposed organisms developed some
tolerance to Pb over time (Arauio et al.. 2019). In a comparative study of adult amphipods Gammarns
fossarum, either freshly collected from the field and exposed to 2.1 |ig Pb/L for 24 hours or chronically
exposed to the same concentration for 10 weeks, there were differences in response to Pb. In the freshly
collected amphipods, both locomotion and respiration were significantly decreased compared with
unexposed organisms, whereas in the chronically exposed amphipods, no statistically significant response
to these endpoints was observed, suggesting that the compensatory response developed over time (Lebrun
and Gismondi. 2020). In another study with G. fossarum and the amphipod Gammarns pulex, a history of
metal exposure did not affect Pb bioaccumulation parameters, as accumulation and elimination
parameters were similar between reference and pre-exposed populations collected from field sites and
exposed to Pb in microcosms (Urien et al.. 2017). Amphipods were exposed to water spiked with an
analytically verified concentration of 10 (ig Pb/L for 7 days, then transferred to mineral water for
depuration for 7 days. The net bioaccumulation of Pb was quantified by subtracting the basal
concentrations of Pb from the total Pb concentration after exposure. There was no interpopulation
variability or difference in the pattern of accumulation or elimination between G. pulex and G. fossarum.
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The peak Pb body concentration was slightly higher in pre-exposed populations relative to the reference
populations for both species.
11.3.2.1.9 Seasonality
In the 2013 Pb ISA, several studies reported seasonal alterations in aquatic plant Pb tissue
concentrations, suggesting that species-dependent seasonal physiological changes may control Pb uptake
in aquatic macrophytes (Section 6.4.9.1) (U.S. EPA. 2013). Several studies published since the 2013 Pb
ISA further describe changes in Pb bioavailability linked to season. In a study examining the interacting
effects of macrophytes and season, metal concentration in small fish inhabiting the phytoplankton-
dominated northern zone of Lake Taihu, China was significantly greater in summer than in small fish
collected from the southern zone of the lake characterized by a high density of macrophytes (Zeng et al..
2012). These differences in metal concentration in small fish collected from the two regions of the lake
disappeared in winter, suggesting that the presence of algae and macrophytes modified trace metal
concentrations during the summer months, resulting in two distinct ecological regions that differed in
their potential for metal exposure. Differences in metal accumulation in larger fish from the two lake
zones varied with season in some tissues, but no significant differences were reported in carnivorous fish.
Chen et al. (2019) quantified seasonal differences in Pb mobility in lake sediments from phytoplankton-
dominated and macrophyte-dominated areas of Lake Taihu. In the phytoplankton-dominated region,
labile and dissolved Pb in sediment was highest in April and July and lowest in October and January. The
opposite pattern was observed for the macrophyte-dominated region. In littoral anoxic sediment, the
periodic drying and rewetting process can increase the bioavailability of Pb to aquatic organisms (Liu et
al.. 2020). Even though high total OC content in the sediment facilitates the formation of anoxic
conditions, periodic drying oxidizes the sediment and leads to sulfide oxidation, which increases the
mobility and bioavailability of Pb because it is less firmly bound to sediment in these conditions.
11.3.2.1.10 Parasites
The combined effects of endoparasites and other stressors, such as metals, modulate uptake and
toxicity to host organisms (Marcogliese and Pietrock. 2011). Multiple studies have reported differences in
Pb accumulation between parasitized and nonparasitized organisms including studies in fish (Brazova et
al.. 2015; Filipovic Marine et al.. 2014; Sures et al.. 2003; Sures and Siddall. 1999) and snails (Mostafa et
al.. 2014). These studies suggest that the effects of parasites on host organisms in the presence of Pb are
complex. In a recent synthesis of parasite-host studies, Pb was accumulated to a higher degree in parasites
than in tissues of host species, and Pb accumulation in infected hosts was consistently lower compared
with uninfected conspecifics (Sures et al.. 2017).
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11.3.2.1.11 Bioturbation/Association with Sediment
Since the 2013 Pb ISA, several studies have examined how the activities of sediment-associated
benthic invertebrates influence Pb transfer to the water column and subsequent bioavailability to other
aquatic organisms. A statistical Random Forest model that took into account riverine invertebrate
community traits such as feeding strategy, respiration and locomotion to predict metal bioaccumulation
from environmental compartments (water column, sediment, suspended particulate matter) showed that
the strongest predictor of metal bioaccumulation in the organisms was the degree to which taxa live in or
directly on sediment (Peter et al.. 2018). In mesocosms with two (Amphipod, Bivalve) or three
(Amphipod, Bivalve, Oligochaete) sediment-associated species combinations, water, and tissue
concentrations of Pb (and other trace elements primarily associated with organic colloids) increased as the
number of bioturbating organisms present increased (Andrade et al.. 2020). One set of experiments
Blankson et al. (2017); Blankson and Klerks (2017. 2016a. 2016b) used oligochaete worm Lumbriculus
variegatus in Pb-spiked mesocosms as a model organism for bioturbation in freshwaters and
demonstrated that concentration of Pb in water column and water turbidity increases with increased
density of sediment organisms. Bioturbation activity was also affected by increasing Pb concentration, a
decline in bioturbation was observed with worms exposed to 681.9 mg/kg and 3396.2 mg/kg (Blankson et
al.. 2017). The amount of Pb transferred to the water column varied with sediment characteristics
(Blankson and Klerks. 2017). For transfer of Pb to the water column, the most important variables were
silt/clay content and sediment pH; Pb bioaccumulation in the worms was influenced by OM in the
sediments and the pH of the porewater. Overall, bioturbation by oligochaetes could bring about the
transport of Pb from sediments to the water column. This means that the presence of these bioturbators
can enhance Pb availability to organisms in the water column and potentially cause toxic effects in
planktonic and nektonic organisms.
11.3.2.1.12 Intraspecific Interactions
Additional research published since the 2013 Pb ISA provides experimental evidence that
interactions among individuals of the same species may affect sensitivity to metals. The influence of
intraspecific competition on Pb (13 to 236 |ig Pb/L) toxicity was explored by Gust et al. (2016) using
single daphnia exposures conducted concurrently with assays of multiple daphnia (proportionally scaled
assays of 20 D. magna per beaker) and at two different feeding regimens (low-feed ration and high-feed
ration). After 14-day exposure to Pb, the LC50 was threefold higher in assays with single daphnia (232
[156-4810] |ig Pb/L) compared with assays with multiple individuals (68 [63-73] |ig Pb/L) at the lower
feeding ration. Similar results were obtained with the higher feeding ration experiment with multiple
daphnia per experimental unit (LC50 = 79 (74-84) |ig Pb/L) and the single-animal treatment
(LC50 = 236 |ng Pb/L) (no 95% confidence interval could be calculated). Moreover, reproduction (neonate
production) decreased with intraspecific competition at 9 and 14 days in both feeding ration groups
compared with assays with single daphnia where no negative effects on reproduction were observed at
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any concentration tested. The authors proposed that individual daphnia modulate their life-history
response in the presence of others of the same species through chemical cues, and this has a modifying
effect on toxicity.
11.3.2.1.13 Predator-Stress and Metal Mixture Effects
Research published since the 2013 Pb ISA tested the effects of multiple stressors on Pb uptake
and toxicity. Predator stress and the presence of other metals affected the accumulation and sensitivity of
the aquatic sowbug (A. aquaticus) to Pb stress (Van Ginneken et al.. 2018). Individual A. aquaticus
collected from a stream in Belgium were placed in a control, 0.0232 |imol/L. (4.8 |ig Pb/L), 0.276 |imol/L
(57 |ig Pb/L) or 3.08 |imol/L (638 |ig Pb/L) solution with two black alder (Alnus glutinosa) leaf discs.
Each Pb treatment and metal mixture (Cu + Pb, Cd + Pb and Cu + Cd + Pb) was crossed with one of two
treatments, either a heterospecific predator cue or conspecific alarm cue. To create the heterospecific
predator cue solution, one damselfly larva (Ccilopteryx splendens) was placed in a container of water for
72 hours. Next, one adult three-spined stickleback (Gasterosteus aculeatus) and the ninespine stickleback
(Pungitius pimgitins) were placed in water for 24 hours. After removing the predators from the water,
equal parts of stimulus water were mixed to create the heterospecific predator cue. To create the
conspecific alarm cue, one A. aquaticus was homogenized in solution. Either water (control) or predator
or the conspecific alarm cue solution was added to the control and Pb-contaminated containers with A.
aquaticus every day for 10 days. Afterward, A. aquaticus Pb concentration, growth rate, feeding rate,
percent active time, survival and respiration rate were recorded. Overall, there were no significant effects
of either heterospecific or conspecific predator cues on Pb accumulation in A. aquaticus, although
respiration rates did increase when exposed to predator cues. Pb accumulation in the isopods was
positively correlated with Pb free-ion activity. There were no significant effects of predator stress on
isopod body burdens. Metal mixture significantly affected Pb accumulation, as the slope of the
relationship between Pb treatment and Pb body burden decreased when Cu and Cd + Cu were added.
Respiration rates were affected by both Pb exposure and predator stress. Differences in respiration rates
between predator-stress and control treatments were greater when isopods had greater Pb body burdens.
Activity levels decreased as Pb body burden increased, but there was no difference between predator
treatments and the interaction between Pb treatment and predator stress. Growth rate (mg/day) was
negatively correlated with Pb free-ion activity in the water but was not found to vary with predator stress
or body burden. Although Pb body burden did not influence feeding rates (mg/mg/day), the Pb body
concentration of A aquaticus exposed to the Pb + Cd mixture had the greatest effect on feeding rate
compared with Pb, Pb + Cu and Pb + Cu + Cd. Finally, activity decreased with increasing Pb body
burden, but was unaffected by predator stress and Pb-metal mixtures.
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11.3.2.2 Uptake and Bioaccumulation in Freshwater Plants and Algae
Studies on bioavailability of Pb in aquatic plants and algae published since the 2013 Pb ISA
continue to support previous findings that plants tend to sequester larger amounts of Pb in roots as
compared with shoots and that there are species-specific differences in uptake of Pb from water and
sediments, as well as compartmentalization of that sequestered Pb (U.S. EPA, 2013, 2006). Further, it
has previously been established that many plants accumulate heavy metals in environments with high
concentrations and are used for phytoremediation at such sites; additional studies on this topic have little
relevance in the current assessment.
Very little new information is available on the bioavailability of Pb in freshwater algae at levels
that are within the concentrations of interest in this ISA (Section 11.1.1). One study contains data on
bioavailability and partitioning between water and sediment correlated with toxic harmful algal blooms
(HABs), which are of concern in many freshwater bodies. This study, conducted in a freshwater reservoir
in Portugal, examined in situ interactions between Pb and Microcystis aeruginosa, a HAB-forming
cyanobacterium found in the United States (Baptista et al., 2014). The metal content of water and
sediments from both the reservoir and an upstream reference site were monitored monthly for 16 months,
during which timeM aeruginosa bloomed twice, firstly forming a scum, and later with colonies scattered
throughout the reservoir. No correlation was found between Pb in the water column and algal blooms.
When blooms occurred, a significant increase of metal levels in the sediment occurred simultaneously
(average Pb concentration was measured at 43.2 mg/kg); however, quantification of the exchangeable
metal fraction during this algal bloom indicated that this Pb was probably not bioavailable. The authors
speculate shallow water depth would have allowed the cells ofM aeruginosa to deposit upon the
sediments rapidly, and the presence of the cyanobacteria in the sediment might have contributed to an
increase in metal content, meaning that algae may be an important biotic compartment for Pb during such
blooms. In three Scottish lakes receiving varying inputs of metals solely from atmospheric deposition
changes in phytoplankton biomass, cellular Pb and the P content of cells were measured simultaneously.
The results showed that algal bloom events in the lakes diluted the mass-specific Pb in the phytoplankton
(Gormlcv-Gallaghcr et al., 2016). As total cellular P increased, there was a corresponding increase in
phytoplankton growth, and the concentration of Pb declined.
In freshwater floating macrophytes, there is also very little new information on the bioavailability
of Pb. These life forms are important because their roots dangle in the water column instead of being
buried in substrate, and thus, Pb uptake occurs solely through the interface with the water column. One
U.S. study examined the uptake and distribution of metals by a floating macrophyte, water lettuce (Pistia
stratiotes L.), in storm water impoundments in Florida (Lu et al„ 2011). Two stormwater impoundment
ponds were divided into two plots, a control without P. stratiotes and one with enough young plants to
initially cover l/20th of the water surface. While the authors stated that water Pb levels were mostly low
(below the Maximum Daily Limits), they did not provide the concentrations. Even at these low
concentrations, reported BCFs of Pb from the water column into plant roots were higher than 104. Lead
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was found inside and adsorbed to plant roots, with approximately 60% of Pb within the root tissue.
Another study by Chen et al. (2019) found that submerged macrophytes in lakes can accumulate Pb,
which is absorbed either from the sediments through roots or from the water by leaves.
Although the U.S. EPA Framework for Metals Risk Assessment states that the latest scientific
data on bioaccumulation do not currently support the use of BCFs and BAFs when applied as generic
threshold criteria for the hazard potential of metals (U.S. EPA, 2007), such metrics are useful to provide
information about the amount of uptake of metals into plants, compartmentalization into different plant
tissues, and differences between species. In a series of field studies undertaken in Sicily, spanning a
gradient of affected wetlands, Pb concentrations in soil, water, and plant tissues of several wetland species
were quantified (Bonanno et al„ 2018; Bonanno and Cirelli, 2017; Bonanno and Vymazal, 2017;
Bonanno et al„ 2017; Bonanno, 2013). These studies affirmed that metal uptake is species-specific
despite similar ecology, anatomy, and life form, and that Pb is mainly compartmentalized in root tissue in
freshwater plants.
11.3.2.3 Uptake and Bioaccumulation in Freshwater Invertebrates
This section expands on the findings from the 1986 Pb AQCD (U.S. EPA, 1986), the
2006 Pb AQCD (U.S. EPA, 2006) and the 2013 Pb ISA (U.S. EPA, 2013) on the bioaccumulation and
sequestration of Pb in freshwater invertebrates. Uptake and subsequent bioaccumulation of Pb varies
greatly between species and across taxa, as characterized in previous U.S. EPA reviews of this metal. In
invertebrates, Pb can be bioaccumulated from multiple sources, including the water column, sediment and
dietary exposures, and factors such as the proportion of bioavailable Pb (Section 11.1.6) lifestage, age and
metabolism can affect the accumulation rate. As reviewed by Wang and Rainbow (2008) and supported
by subsequent studies, there are considerable differences between species in the amount of Pb taken up
from the environment and in the levels of Pb retained in the organism.
Uptake studies generally show that aquatic invertebrates accumulate Pb from water in a
concentration-dependent manner and may reach an equilibrium depending on the organism's ability to
eliminate or store Pb. In a study reviewed in the 2013 Pb ISA, the tissue concentration of Pb in adult
Eastern Elliptio mussel (Elliptio complcmata) increased for the first 14 days in an aqueous exposure at an
exposure-dependent rate then did not change significantly for the remainder of the 28-day exposure
(Mosher et al., 2012b) In another study with the same species conducted after the 2013 Pb ISA, Pb was
measured in hemolymph every 7 days during a 28-day exposure, and distinct patterns of response were
observed with Pb concentration. At the lowest concentrations (< 6 |ig Pb/L), Pb gradually increased in the
hemolymph but did not exceed the exposure concentration, at midrange concentration (up to 66 |ig Pb/L),
the mussels appeared to regulate Pb by day 14, whereas at the highest concentration tested (251 |ig Pb/L),
Pb in hemolymph increased throughout the exposure period (Mosher et al„ 2012a). Pb in tissue was
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highly correlated with the exposure concentration at the end of the experiment. The lowest exposure
concentration of 0.9 |ig Pb/L resulted in an average tissue concentration of 1.5 |ig Pb/g dry weight.
Studies in the 2006 Pb AQCD and 2013 Pb ISA generally showed that the tissue distribution of
Pb in aqueous exposures of freshwater invertebrates is primarily sequestered in the gills, hepatopancreas
and muscle. Recent short-term (3-4-hour) aqueous uptake studies with juvenile snail L. stagnctlis showed
no significant difference in Pb accumulation among foot, mantle, digestive tract and remaining soft
tissues, suggesting uptake occurred directly across the skin (Cremazy et al., 2019). L. stagnctlis was
previously identified as one of the aquatic invertebrates most sensitive to Pb exposure (Grose 11 and Brix,
2009; Grosell et al.. 2006b).
There is some evidence to suggest patterns of tissue distribution differ when uptake of Pb is from
sediment. In 28-day exposure to Pb-spiked sediments (205 ± 9 and 419 ± 16 mg/kg dry mass) the
freshwater bivalve Hyridellct australis accumulated Pb in both the low (2.2 ± 0.2 mg/kg dry mass) and
high treatments (4.2 ± 0.1 mg/kg dry mass) in the order labial palps>mantle>gill>visceral mass>muscle
(Marasinghc Wadigc et al„ 2014). Labial palps accumulated significantly more Pb than other tissues,
consistent with the sediment-burrowing activities of this species. After 28-days, 83%—91% of the
accumulated Pb in hepatopancreas of the bivalves was in the biologically detoxified fraction, primarily
sequestered in MRG. Concurrently, the relative proportion of Pb sequestered in the metallothionein-like
protein fraction (13% to 32%) decreased with Pb exposure. The biologically active metal fraction
significantly increased with increased Pb exposure, and the highest percentage of Pb was associated with
the mitochondrial fraction.
The 2006 Pb AQCD recognized the potential importance of the dietary uptake pathway as a
source of Pb exposure for invertebrates. Additionally, several studies reviewed in the 2013 Pb ISA
quantified water versus dietary uptake of Pb in aquatic invertebrates (Komjarova and Blust, 2009;
Borgmann et al., 2007; Besser et al., 2005). Since the 2013 Pb ISA, the relative importance of dietary
versus aqueous uptake pathways has been further discerned for some biota. Camusso et al. (2012) applied
a biologically based Biodynamic Model to previously published data and additional unpublished data on
uptake of trace metals in L. variegatus from field-collected sediments to assess the main uptake route in
this sediment-dwelling organism. The modeled data suggest that for Pb, both free dissolved concentration
in porewater and dietary uptake contributed to body burden, and the amount of Pb taken up in the gut
appears to be controlled by how tightly Pb is bound to sediment. In D. magna fed under two different
dietary regimens (regular diet = 3/105 Raphidocellis subcapitata algal cells/mL; restricted diet = half
algae concentration), Pb uptake from water was gradual in individuals with restricted food intake and
faster under regular feeding, suggesting that a portion of Pb uptake occurred via diet (Araujo et al.,
2019). Hadji et al. (2016) used a series of microcosms in which the amphipod G. pulex was exposed to
Pb for 6 days in the water column only (0.36, 0.71. 3.62, 6.75 (ig Pb/L) or water column (0.31, 0.57,
3.07, 5.02 |ig Pb/L) with access to food (poplar leaves Poplns nigra pretreated for 1 week in Pb
concentrations ranging from 0.5 to 10 |ig Pb/L). In the water-column-only microcosms, Pb-treated poplar
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leaves were present but were enclosed in mesh bags so that the gammarids could not feed. At the end of
the study, Pb was significantly higher in amphipods with access to Pb-contaminated leaves than in
amphipods exposed to Pb via the water column alone. The dietary contribution ranged from 29% to 31%
in the tested concentrations. In an 8-day depuration period, there were no significant differences in
elimination regardless of exposure route.
Few studies have assessed the relationships between Pb speciation, water chemistry and
biouptake in aquatic invertebrates in situ. In aquatic insect taxa (Leuctrci spp., Simuliidae, Rhithrogena
spp, Perlodidae) sampled from 36 headwater streams in the Lake District of England, pH was the
prevalent factor influencing Pb uptake, and there were statistically significant correlations between total
dissolved Pb in stream water and insect body burdens (De Jongc et al„ 2014). For prediction of observed
body burdens, Windermere Humic Aqueous Model modeling of stream chemistry and Pb chemical
speciation that took into account competition among cations for uptake in biota resulted in a better model
fit than "metal accumulation as a function of total dissolved metal levels or the free ion alonc"(Dc Jonge
etal.. 2014).
11.3.2.4 Uptake and Bioaccumulation in Freshwater Vertebrates
In freshwater vertebrates, Pb uptake in tissues generally increases with increasing concentration
of Pb in exposure media (U.S. EPA, 2013); recent studies continue to support these observations.
Evidence in the 2013 Pb ISA supported the 2006 AQCD conclusions that the gill is a major site of Pb
uptake in fish and that there are species differences in the rate of Pb accumulation and distribution of Pb
within the organism. In dietary studies reviewed in the 2013 Pb ISA, the anterior intestine was identified
as a target of Pb in fish. New uptake studies continue to show distinct patterns of Pb tissue distribution in
water versus dietary exposures. As reviewed in Lee et al. (2019), some studies in fish reported higher
rates of Pb accumulation in gill tissues from waterborne exposure compared with dietary exposure. Pb
typically accumulates in metabolically active organs including kidney, liver, and intestine in both aqueous
and dietary exposure.
In a study designed to investigate the relative influence of waterborne and dietary Pb on
accumulation by rainbow trout (O. mykiss), juvenile trout were exposed to Pb ( 8.5, 20, 60 or
110 |Lig Pb/L), for 7 weeks via waterborne Pb only, dietary Pb only in the form of live prey (worms L.
variegatus pre-exposed for 28-days to the same concentration of Pb as the fish) or simultaneously to
waterborne and dietary Pb (Alsop et al., 2016). Accumulation of Pb in fish was significantly higher via
the waterborne exposure pathway compared with dietary exposure in all tissues except in the gut, which
accumulated similar amounts of Pb regardless of the exposure route. When fish were exposed to Pb from
both water and their diet, whole-body Pb was reduced up to 61% at 110 |Lig Pb/L, and Pb accumulation
was significantly reduced at a threshold of ~50 |ig Pb/L, with significantly lower concentrations in the
liver and carcass but not the gill or gut. The authors noted that Pb may have altered the nutrient quality of
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the prey; carbohydrates and lipid levels in the worms were significantly decreased even at the lowest Pb
concentration.
11.3.2.5 Uptake and Bioaccumulation Through Food Web
In the 2006 Pb AQCD (U.S. EPA, 2006) and the 2013 Pb ISA (U.S. EPA, 2013), transference of
Pb through the food web was generally found to be low, with lower Pb accumulation at higher trophic
levels; however, some studies found bioaccumulation of Pb at higher trophic levels. Recent evidence
supporting little bioaccumulation through freshwater food webs is reviewed here.
In a review published since the 2013 Pb ISA, Cardwell et al. (2013) compiled laboratory and field
studies published prior to the 2013 Pb ISA to examine the transfer of Pb and other heavy metals through
aquatic food webs. The concentrations of Pb decreased with increasing trophic position in food web
studies examining trophic transfer between phytoplankton, cladocera and fish. In most of the field studies
reviewed, no evidence was found for biomagnification of Pb across trophic levels in freshwater systems.
Specifically, 17 studies examined trophic transfer of heavy metals through aquatic lake or stream food
webs; while 10 of these studies found no evidence of Pb biomagnification, one study found possible
evidence, and six studies did not examine Pb or did not present data on Pb. More recent studies are
presented below.
In a high-elevation lake in the Alps, Pastorino et al. (2020b) examined the accumulation of heavy
metals, including Pb, in sediment, chironomids, and fish. Surface sediment, benthic macroinvertebrates,
and fish were sampled from a glacial-origin lake, Dimon Lake, in Northeast Italy. While there is only a
single fish species in this lake, i.e., the European bullhead (Cottiis gobio), the benthic macroinvertebrate
community consists of midges (Diptera Chironomidae), worms (Oligochaeta), and leeches (Hirudinea).
The only prey found in the stomachs of C. gobio was Diptera Chironomidae, and therefore only these
specimens were used for trace-element analysis. Surface sediment Pb was 109.6 ±1.2 mg Pb /kg, whole-
body Diptera Chironomidae Pb concentration was 49 ± 0.5 mg Pb/kg, and Pb concentration in C. gobio
was 0.06 ± 0.03 mg Pb/kg in the muscle and 0.03 ± 0.4 mg Pb/kg in liver. The BAF and trophic transfer
factor (TTF) in Diptera Chironomidae and C. gobio muscle and liver samples were less than 1.0 for Pb,
indicating biodilution. The BAF in Diptera Chironomidae was 0.45 and the BAF in fish muscle and liver
was 0.0005 and 0.003, respectively. The TTF in C. gobio was 0.002 in muscle and 0.007 in liver. In a
similar study, Pastorino et al. (2020a) examined BAFs for all the benthic macroinvertebrates from Dimon
Lake (Chironomidae, Oligochaeta and Hirudinea) and from another nearby lake, Balma Lake
(Chironomidae, Oligochaeta). In this analysis, Dimon Lake surface sediment was 110 ± 1.1 mg Pb/kg
(mean ± S.D.), while Balma Lake had considerably less Pb (41 ± 1.2 mg/kb Pb). In addition to lower Pb
concentration in the surface sediments, Balma Lake had a lower pH (mean ± S.D.; summer: 6.70 ± 0.34;
autumn: 7.64 ± 0.09) than Dimon Lake (summer: 8.77 ± 0.12; autumn: 9.44 ± 0.05). The lower pH was a
result of Balma Lake's granite bedrock whereas Dimon Lake covers volcanic rock and sandstone. No
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correlation was found between the sediment trace-element concentrations and the benthic
macroinvertebrates. BAFs were calculated using the mean Pb sediment concentration from each lake
across the summer and the fall. In this study, BAFs for Dimon Lake Chironomidae were similar to results
found in Pastorino et al. (2020b) for Chironomidae at 0.45. Additionally, Dimon Lake BAFs were 0.42
for Oligochaeta and 0.1 for Hirudinea, suggesting biodilution. In Balma Lake, however, BAFs were
above 1.0, suggesting bioaccumulation for the benthic macroinvertebrate community (1.61 for
Chironomidae and 1.66 for Oligochaeta).
Some studies use stable-isotope analysis to characterize trophic position in a food web. Using
stable isotopes, Pb accumulation was found to decrease with increasing trophic level in Korean wetlands
(Kim and Kim. 2016). The Upo wetlands consist of four smaller wetlands (Upo, Mokpo, Sajipo, and
Jokjibul), which have different water inflow sources and consequently abiotic condition and biotic
communities. Sediment and biota (primary producers: water caltrop [Trcipci japonica], primary
consumers: leaf beetle [Galerucella nipponensis\ and secondary consumers: water strider [Gerris sp.] and
wolf spider [Arctosct sp.]) were collected and characterized for metal content (Pb, Cd, Cu, and Zn).
Afterward, 813C and 815N isotopes were used to characterize the food web. Sediment Pb concentrations
ranged from approximately 35 to 50 mg Pb/kg and differed significantly among sites. In general, the plant
and leaf beetle had lower 813N and 815N signatures than water striders and wolf spiders. Concentrations
of Pb in the leaves of T. japonica were the highest compared with the other organisms analyzed at all
sites. Pb concentrations in G. nipponensis were significantly lower than those in T. japonica. Pb
accumulation in the secondary consumer Arctosa sp. was lower than Pb accumulation in Gerris sp.
Overall, Pb concentrations decreased significantly as trophic level increased (plant
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contaminant exposure or the reach-specific spider mean metal concentration, divided by the toxic
threshold for each study reach. Pb chronic risk quotients calculated for the Emory River study area ranged
across species, with the highest risk quotients found for 1 and 12-day Chickadee nestlings (Poecile spp.)
(range: 0.81-1.52; percentage of diet consisting of spiders: 25.0%), Eastern Bluebird 2-day nestlings
(Sialic/ sialis) (range: 0.81-0 1.21; percentage of diet consisting of spiders: 30.9%), and Red-cockaded
Woodpecker 9-12-day nestlings (Picoides borealis) (range: 0.80-1.20, percentage of diet consisting of
spiders: 60%). All Pb acute risk quotients reported were 0.00. Chronic spider-based avian wildlife values
for adult and nestling birds ranged from 0.03 mg Pb/kg for 1-day nestlings for Poecile spp. to
I.347 mg Pb/kg for Setophaga discolor (prairie warbler) 12-day nestlings.
In another example of aquatic insect transfer of Pb to the surrounding environment, Fletcher et al.
(2022) found that 80%-95% of Pb in dragonfly species was shed with emergence. Ten dragonfly species
were collected from a constructed wetland at the Savannah River Site, a National Environmental
Research Park in South Carolina, United States, where materials for nuclear weapons are produced.
Although sediment and freshwater concentrations were not reported in this study, average Pb
concentrations in the shed exuviae of 10 dragonfly species (Brachymesia gravida, Libelhda auripennis,
Libellula luctiiosa, Orthemis ferruginea, Plathemis lydia, Pachydiplax longipennis, Perithemis tenera,
Pantala flavescens, Pantala hvmenaea, and Tramea lacerata) ranged from 2.94 to 10.7 mg Pb/kg, which
was significantly higher than Pb concentrations in the tenerals, or the freshly molted adult insect
(< 0.4 mg Pb/kg), suggesting that Pb in the exuviae was 17-96 times higher than the concentrations in the
teneral.
New observational studies and literature reviews since the 2013 Pb ISA (U.S. EPA. 2013)
generally confirm that many freshwater food webs exhibit reduced accumulation of Pb in higher trophic
levels (Pastorino et al.. 2020b; Kim and Kim. 2016; Cardwell et al.. 2013). although one study reported
the bioaccumulation of Pb (Pastorino et al.. 2020a). Additional studies demonstrated that Pb can transfer
between aquatic food webs and terrestrial ecosystems via aquatic insect emergence and predation by and
of riparian spiders (Fletcher et al.. 2022; Kraus et al.. 2021; Beaubien et al.. 2020).
II.3.3 Environmental Concentrations of Pb in Freshwater Biota and
Ecosystems in the United States at Different Locations and Over Time
Few U.S. studies have examined national or regional-scale trends of Pb in freshwater biota. The
1986 AQCD reported the results of Lowe et al. (1985). a nation-wide survey of metal concentrations in
fish from 1979 to 1981. At 112 monitoring stations, they found an average (geometric mean) of
0.19 |ig Pb/g wet weight for the period 1978 to 1979 and 0.17 |ig Pb/g wet weight for 1980 to 1981 (U.S.
EPA. 1986). In the 2006 AQCD, a representative median and a range of Pb concentrations were reported
in surface waters (median 0.50 |ig Pb/L, range 0.04 to 30 |ig Pb/L), sediments (median 28 mg Pb/kg dry
weight, range 0.5 to 12,000 mg Pb/kg dry weight) and fish tissues (geometric mean 0.54 mg Pb/kg dry
weight, range 0.08 to 23 mg Pb/kg dry weight [whole body]) in the United States based on a synthesis of
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National Ambient Water Quality Assessment data (U.S. EPA, 2006). The 2013 Pb ISA reported survey
results from the Western Area Contaminants Assessment Project (2002-2007), which included the
concentration of Pb in fish tissue (0.0033 [fillet] to 0.97 [liver] mg Pb/kg [dry weight]) from a set of
national parks in the western United States (Blctt. 2010: U.S. EPA, 2008b). No recent studies examining
spatial or temporal trends in Pb concentration in freshwater fish or invertebrates from locations across the
United States were identified in this ISA. Many individual studies report Pb concentrations in aquatic
ecosystems and biota from specific sites across the United States; compilation of those data is outside the
scope of this ISA. Pb concentrations in water, sediment and other environmental media are available in
Section 11.1.3 and summarized in Table 11-1.
Since the 2013 Pb ISA, a few regional-scale studies, including a study in Canada, have assessed
trends in Pb concentrations in vegetation (peat bogs) or the water column. Peat bogs deposit and preserve
stable layers of accumulated moss and other plant material that can be used to reconstruct a record of
spatial and temporal distribution patterns of air Pb concentrations. Six peat cores collected in 2013 and
2014 in northern Alberta, Canada Shotyk et al. (2016) record the rates of atmospheric Pb deposition dated
from 1910 to 2014 using 21"Pb and 14C dating in models, linking sample depth to age. Peak accumulation
rates were observed between the years 1950 and 2000 in each sample, and overall decreasing rates of Pb
accumulation were observed from 1980. Although this study was not in the United States, decreased Pb
accumulation rates coincided with the introduction of unleaded gasoline in the United States and Canada
in the mid-1970s and nearby potential point sources of Pb air pollution (industrial development including
bitumen mines and upgraders) are not attributed to the increase in Pb accumulation. The uppermost, most
recent, peat layers show near-zero modern atmospheric Pb deposition in the Alberta peat bogs.
In a 2015-2017 water quality survey of four Tennessee headwater Appalachian streams Olson et
al. (2019), the maximum observed Pb concentration and sole detectable measurement of this metal was
less than 1 |ig Pb/L. Reported mean concentration values at each site were less than the minimum
detection limit of 0.28 |ig Pb/L. These observations from remote streams without upstream anthropogenic
Pb sources suggest that long-range atmospheric deposition is not a major source of Pb contamination to
this region. Limited evidence from regional studies of temporal trends in freshwater aquatic ecosystems
published since the 2013 Pb ISA suggest that modern atmospheric deposition of Pb is not a major
contributor to Pb concentrations in freshwater aquatic biota and ecosystems in remote locations.
11.3.4 Effects of Pb in Freshwater Systems
This section focuses on studies of the biological effects of Pb on freshwater plants and algae,
microbes, invertebrates, and vertebrates published since the 2013 Pb ISA. The biological effects of Pb in
the 2013 Pb ISA and in this appendix are generally presented in increasing complex levels of biological
organization from suborganismal responses (i.e., enzyme activities, changes in blood parameters) to
endpoints relevant to the population level and higher (growth, reproduction, and survival) up to effects on
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ecological communities and ecosystems. Exposure-response studies that report toxicological dose
descriptors (e.g., LC50, EC50, lowest observed adverse effect level [LOAEL]) for effects on growth,
reproduction or survival endpoints are reported in Section 11.3.5.
11.3.4.1 Effects on Freshwater Microbes
The effects of Pb on microbial communities in freshwater ecosystems were not reviewed in detail
in the 2013 Pb ISA (U.S. EPA, 2013), except for a report that Pb could alter bacterial infection in the fish
Charmapunctatiis (Pathak and Gopal, 2009). In the 2006 Pb AQCD (U.S. EPA, 2006), it was reported
that Pb could adsorb to biofilms, depending on pH, water hardness, polarity of matter, and amount of Fe
or Mn in the water and that methylation by microbes may result in Pb remobilization in aquatic
ecosystems; however, few studies directly report effects on microbes from Pb exposure. Since the 2013
Pb ISA (U.S. EPA, 2013), several experimental and observational studies have examined the relationship
between Pb concentration in the sediment and effects on freshwater microbes, as summarized below.
Several of these studies report negative relationships between sediment Pb concentration and microbial
abundance or community structure, while some report no relationship or positive associations.
In a study from the United States, porewater and sediment Pb concentrations were negatively
correlated with bacterial RNA abundance, but not diversity or richness in Lake DePue, Illinois (Gough
and Stahl, 2011). Lake DePue is a shallow lake on the Illinois River located near a U.S. EPA Superfund
Site (the DePue/New Jersey Zinc/Mobil Chemical Site). Although the Zn smelting facility and phosphate
fertilizer plant are no longer operational, Lake DePue has received metal-contaminated sediments from
this site for over 80 years. Sediment Pb concentration in the lake was on average 180 mg Pb/kg (range:
68.6 and 541 mg Pb/kg). Porewater and sediment Pb were correlated with a low abundance of archaeal,
bacterial, and eukaryotic terminal restriction fragments (TRF). Overall, there were some differences in
overall community structure with regard to metal contamination observed using terminal restriction
fragment length polymorphism analysis of 16S rRNA genes, although variation in bacterial diversity,
richness and composition across a metal gradient was not detected. In a follow-up study using the same
samples, Kang et al. (2013) further explored the bacterial communities using a different approach, a
function gene microarray (GeoChip). Overall, the diversity of functional gene variants was similar across
all five sites, suggesting that heavy-metal concentrations in the sediments did not significantly affect
bacterial community structure; however, some individual gene categories were correlated with certain
porewater metal concentration, including Pb. Using a CCA, Pb, Zn, and Cd were all found as important
predictors for sulfate-reducing bacteria communities. Although significant correlations with Pb existed,
functional gene variants had similar relationships with other porewater metal concentrations, including
As, Cd, Cr, and Zn.
In another U.S. study, observational evidence suggests that the exchangeable Pb fraction
decreases microbial community diversity, while oxyhydroxide Pb concentration was correlated with an
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increase in diversity in the mining district of Lake Coeur d'Alene. Idaho (Mobcrlv et al.. 2016). Pb
concentrations in the sediment were high in the lake, ranging from 1,540 to 3,422 mg Pb/kg, while the St.
Joe River delta reference site sediment Pb concentration was 29 mg Pb/kg. More than 70% of the Pb was
associated with the exchangeable/carbonate phase, which is thought to the be most bioavailable phase. Pb
in the exchangeable/carbonate fraction was negatively correlated with the abundance of Aquificae and
Synergistes and positively correlated with candidate phylum LD1PA abundance (a phylum without many
cultured representatives); furthermore, this pattern is similar for Fe and Mn oxyhydroxides, as Pb
exchangeable/carbonate concentrations are highly correlated. Bacteroidetes OTU abundance was
negatively correlated with Pb-exchangeable/carbonate and positively correlated with Pb-(oxy)hydroxide.
These results suggest that the phase of Pb is integral in determining the relationship between Pb
concentration in the sediment and microbial communities, as seasonal changes in Pb speciation could
affect microbial diversity.
To understand how heterotrophic bacteria in river sediments are affected by Pb, sediments were
collected from sites along three tributaries of the Nagara River in Japan, varying in land use types
(agricultural, industrial, or forested) and contamination (Du et al.. 2018). Overall, Pb (100, 1.000, and
10.000 |ig/L) did not have significant effects on heterotrophic bacteria density, activity, and community
structure after 30 days of a sequencing batch incubation experiment.
In contrast, the Pb enrichment factor, along with other heavy metals, was found to influence
bacterial community structure in the Poyang Lake river system, China (Zhang et al.. 2018). Mean Pb
concentration in the sediments from five rivers ranged from 29.52 to 40.06 mg Pb/kg. The Pb enrichment
factor, which takes into account Fe as the normalizer element, along with the As and Cd enrichment
factors, pH, OC, and degree of contamination were the main variables affecting bacterial community
structure (redundancy analysis). The Pb enrichment factor, as well as Cd enrichment factor and the degree
of contamination, was strongly associated with higher abundances of Acidobacteria, suggesting tolerance
of the phyla.
In another study in freshwater systems in China, Pb concentration in the sediment was found to
negatively correlate with the relative abundance of major bacterial groups, but not with bacterial diversity
(Li et al.. 2020). Pb concentration in the sediment was 17.3 ± 7.3 mg Pb/kg (mean ± S.D.) and ranged
from 1.9 to 25.4 mg Pb/kg across 12 sites in Huangjinxia Reservoir in Shaanxi Province, China. Sediment
Pb concentration was highly correlated with Cr, Zn, and Ni but not significantly correlated with microbial
diversity indices (ACE, Chaol, Shannon, and Simpson's index). However, Pb sediment concentration
was significantly negatively correlated with the relative abundance of Bacteroidota, Nitrospirota, and
Verrucomicrobia and positively correlated with the relative abundance of Chloroflexi..
Variation in bacterial community composition along an elevation gradient in Yangtze River,
China, was driven by OM, elevation, urbanization, and Pb concentration (Zhang et al.. 2020). Pb
concentration in the sediment ranged from 14.40 ± 0.80 mg Pb/kg to 87.01 ± 8.00 mg Pb/kg. Elevation
(meters above sea level) was negatively correlated with Pb concentration in the sediment as were many
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other physicochemical parameters and metal concentrations. OM was the most significant variable,
followed by elevation (10.4%), urbanization rate (9.0%) and Pb (9.5%). Bacterial community structure
between 50 and 400 masl was most correlated with Pb, and below 50 masl community structure was most
correlated with urbanization rate. Above 400 masl, Pb concentration and OTU abundance were
significantly correlated, while the correlations between Pb and the Shannon index and evenness were not
significant. Below 400 masl, the opposite pattern emerged: the relationship between Pb and OTU
abundance was negative and the relationships between Pb and the Shannon index and evenness were
nonsignificant. Finally, Pb concentration was positively correlated with the abundance of certain bacterial
genera, negatively correlated with others, and not correlated with most dominant bacteria taxa.
In summary, since the 2013 Pb ISA (U.S. EPA. 2013). several observational and experimental
studies examining the effects of Pb concentrations in freshwater sediment and porewater found negative
associations with bacterial or archaeal abundance, but not diversity (Li et al.. 2020; Kang et al.. 2013;
Gough and Stahl. 2011). while others found mixed associations between Pb and microbial diversity
(Moberlv et al.. 2016) or no relationship (Du et al.. 2018).
11.3.4.2 Effects on Freshwater Plants and Algae
The toxicity of Pb to freshwater algae and plants has been recognized in earlier U.S. EPA reviews
of the metal and the findings are briefly summarized here. In the 1977 Pb AQCD, differences in
sensitivity to Pb among different species of algae were observed, and concentrations of Pb within the
algae varied among genera and within a genus (U.S. EPA. 1977). The 1986 Pb AQCD (U.S. EPA. 1986)
reported that some algal species (e.g., Scenedesmus sp.) were found to exhibit physiological changes
when exposed to high Pb concentrations in situ. Effects of Pb on algae reported in the 2006 Pb AQCD
included decreased growth, deformation, and disintegration of algae cells, and blocking of the pathways
that lead to pigment synthesis, thus affecting photosynthesis. Most studies on effects of Pb in freshwater
algal species reviewed in the 2013 Pb ISA and the AQCDs were conducted with nominal media
exposures and effect concentrations greatly exceeded Pb reported in surface water. In studies in which Pb
was quantified, effect concentrations for growth (EC50) for freshwater algae and macrophytes were much
higher than currently reported environmental Pb. Growth endpoints in freshwater algae reviewed in the
2013 Pb ISA included significant inhibition of chlorophyll a content at 210 (ig Pb/L and higher in Wolffict
ctrrhizct (Piotrowska et al.. 2010). An increase in biomass was reported in L. minor exposed to 100 or
200 |ig Pb/L, with inhibition observed at higher concentrations (Dirilgen. 2011). There were also
numerous studies conducted at nominal Pb concentration that reported effects on enzyme activities and
protein content in freshwater aquatic plant species. Exposure-response relationships in which increasing
concentrations of Pb lead to increasing effects were consistently observed for freshwater aquatic plants. In
the 2013 Pb ISA, the body of evidence was sufficient to conclude there were likely to be causal
relationships between Pb exposure and freshwater plant physiological stress and between Pb exposure and
reduced freshwater plant growth. The body of evidence was inadequate to conclude there are causal
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relationships between Pb exposure and freshwater plant reproduction and between Pb exposure and
freshwater plant survival.
New information on freshwater algae since the 2013 Pb ISA addresses the deficit of analytically
verified chronic toxicity data for these organisms. De Schamphelaere et al. (2014) conducted 72-hour
bioassays in standard test media to assess growth rate in three commonly tested algal species; P.
subcapitata, C. kesslerii, and C. reinhardtii. P. subcapitata was the most sensitive, with
EC50 = 83.9 |ig Pb/L, EC20 = 45.7 |ig Pb/ and EC10 = 32.0 |ig Pb/L based on filtered Pb concentration.
Furthermore, in subsequent tests with P. subcapitata at varying pH, 72-hour EC50 decreased from 72.0 |ig
filtered Pb/L at pH 6.0 to 20.5 |ig filtered Pb/L at pH 7.6. The authors noted that this species exhibited
greater sensitivity to Pb than two of the most chronically Pb-sensitive aquatic invertebrates (the
crustacean C. dubia and the snail L. stagnalis) at pH > 7.4 based on model-predicted chronic EC50 values.
Additionally, new information on Pb effects on the emergent freshwater macrophyte, the common
reed (Phragmites aiistralis), shows an alteration in growth form and propagation strategy under Pb
exposure. In a phytotron experiment, reed plants were exposed to five Pb levels in sediment (measured
5.9 ± 0.2, 304 ± 4.38, 508 ± 7.89, 1513 ± 37.28, 3020 ± 120.41 mg Pb/kg) (Zhang et al.. 2015a). In
addition to decreases in total biomass, photosynthesis and rhizome growth, the addition of Pb caused a
significant alteration in growth form. The numbers of axillary shoot buds and daughter apical rhizome
shoots were increased by Pb addition at the highest concentrations, and the bulk (80%) of daughter shoots
were from daughter axillary shoots. This clonal propagation strategy of increased formation and output of
axillary shoot buds, called the phalanx pattern, is an adaptive response to maintain population stability at
the lowest energetic cost. This same growth pattern alteration was also found in an additional study on the
effects of Pb and drought in P. aiistralis by the authors (Zhang et al.. 2015b). but clonal modular growth
and reproductive ability were significantly inhibited by the interaction between drought and Pb. These
propagation effects would cause a decline in P. aiistralis populations in a dry environment under Pb
pollution.
In summary, information published since the 2013 Pb ISA does not substantially change what
was previously known about Pb effects on freshwater plants and algae. A few new studies assessed the
sensitivity of freshwater algal growth to Pb exposure and found a significantly negative effect in certain
species. New information on Pb effects on common reed (P. aiistralis) shows significant decreases in
total biomass, photosynthesis, and rhizome growth as well as alterations in growth form and propagation
strategy under Pb exposure. The growth and reproductive ability of common reed have also been shown
to be significantly inhibited by an interaction between Pb exposure and drought, which may have
implications for future drought events. There is still little information on the relationships between Pb
exposure and freshwater plant or algal survival, particularly at exposure levels below the thresholds used
in this ISA.
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11.3.4.3 Effects on Freshwater Invertebrates
Freshwater aquatic invertebrates are generally more sensitive to Pb exposure than other taxa.
Controlled studies at concentrations near the upper range of representative concentrations of Pb available
from surveys of U.S. surface waters (median: 0.50 |ig Pb/L; range 0.04 to 30 |ig Pb/L, 95th percentile 1.1
Hg Pb/L) (U.S. EPA, 2006) (Table 11-1) reviewed in the 1986 AQCD, the 2006 Pb AQCD and the 2013
Pb ISA provide evidence for the effects of Pb on reproduction, growth and survival in sensitive
freshwater invertebrates. Freshwater invertebrate taxa that exhibit sensitivity to Pb include some species
of gastropods, amphipods, cladocerans and rotifers, although the toxicity of Pb is highly dependent upon
water quality variables such as DOC, hardness, and pH. Key studies reported in the 1986 AQCD include
increased mortality as low as 19 |ig Pb/L for the snail Lvmnaea pcilustris (Borgmann et al., 1978) and
reproductive impairment at 30 |ig Pb/L (nominal values) for Daphnia sp. (Biesinger and Christensen,
1972). In a 42-day chronic study reviewed in the 2006 Pb AQCD, the LOEC for reproduction was
3.5 (ig Pb/L in the amphipod H. ciztecci receiving both waterborne and dietary Pb (Besser et al„ 2005).
In the 2013 Pb ISA, additional studies provided evidence for Pb effects on freshwater
invertebrates at low |ig Pb/L concentration. The growth of juvenile freshwater snails (L. stagnctlis) was
inhibited at an EC20 of <4 |ig Pb/L (Grosell and Brix, 2009; Grosell et al., 2006b). In fatmucket mussel, L.
siliquoidea juveniles, a chronic value (geometric mean of no-observed-effect concentration [NOEC] and
LOEC) of 10 |ig Pb/L was obtained following 28-day exposures (Wang et al„ 2010). In a 7-day exposure
of the cladoceran C. dubict to 50 to 500 |ig Pb/L, increased DOC led to an increase in mean EC50 for
reproduction ranging from approximately 25 |ig Pb/L to >500 |ig Pb/L (Mager et al., 201 la). The 48-hour
LC50 values for the cladoceran C. dubict tested in eight natural waters across the United States varied from
29 to 1,180 |ig Pb/L and were correlated with DOC (Esbaugh et al., 2011). The freshwater rotifer E.
dilatata 48-hour LC50 was 35 |ig Pb/L using neonates hatched from asexual eggs (Arias-Almeida and
Rico-Martinez, 2011). The EC20 for reduced growth and emergence of the midge C. dilutus was reported
to be 28 |ig Pb/L, observed in a 55-day exposure study, while the same species had a 96-hour LC50 of
3,323 |ig Pb/L (Mebane et al., 2008) The EC10 for molting in the mayfly B. tricaudatus was 37 |ig Pb/L
(Mebane et al., 2008). These studies provided evidence in the 2013 Pb ISA supporting determinations of
causal relationships between Pb exposure and growth, reproductive effects, and survival in freshwater
invertebrates (Table 11-4).
11.3.4.3.1 Suborganism-Level Response
The key studies described above from the 2013 Pb ISA and earlier AQCDs report effects on
reproduction, growth, and survival in freshwater invertebrates. Additional endpoints for Pb toxicity in
aquatic invertebrates considered in the 2013 Pb ISA and previous AQCDs included suborganism-level
effects such as enzyme function and oxidative stress. These suborganism-level effects were considered
together in the 2013 Pb ISA as "physiological stress" and the body of evidence was sufficient to conclude
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that there is a likely to be causal relationship between Pb exposure and altered response. Although stress
responses are correlated with Pb exposure, they are nonspecific and may be altered with exposure to any
number of environmental stressors. An additional suborganism-level endpoint in the 2013 Pb ISA was
"hematological effects," which included changes to ALAD expression or the hematopoietic system
associated with Pb exposure. For this endpoint, the body of evidence was sufficient to conclude that there
is a likely to be causal relationship between Pb exposure and hematological effects in freshwater
invertebrates in the 2013 Pb ISA. These suborganism-level responses may serve as biomarkers for effects
at the organism level and higher; however, only a subset of studies that quantified response at the
suborganismal level concurrently assessed effects on growth, reproduction, development, or survival.
Only a few of the many studies identified in the literature search on suborganism-level response to Pb
exposure in freshwater invertebrates were conducted in the low |ig Pb/L range and hence met the criteria
for inclusion in the ISA.
Recent literature supports the previous evidence for Pb effects on enzymes and antioxidant
activity in freshwater invertebrates. New studies on physiological stress endpoints include changes in the
activities of antioxidant defense enzymes with aqueous exposure to Pb. SOD and GPx activities were
significantly reduced, and MDA levels were significantly increased in juvenile Oriental river prawn
(Macrobrachium nipponense) exposed to 25 |ig Pb/L for 60 days. CAT activity in the hepatopancreas
increased at 12 (ig Pb/L and decreased in the 25 |ig Pb/L treatment (Ding et al., 2019). In the same study,
reductions in weight gain and specific growth rate were observed in prawns exposed to 25 |ig Pb/L in
chronic 60-day exposure tests. No growth effects were observed in prawns at 12 |ig Pb/L (see
Section 11.3.5).
Physiological stress in freshwater invertebrates was also assessed during sediment exposure to
Pb. Exposure of larval midge Chironomus riparius to Pb-spiked sediment (132 mg Pb/kg dry weight and
505.5 mg Pb/kg dry weight) for 16 days resulted in an antioxidant response (increase in metallothionein)
and cellular damage (increase in MDA) (Arambourou et al., 2013). There was no significant change to
protein concentration, lipid was depleted while glycogen increased with increasing Pb in the sediment. In
the same organisms, Pb exposure via sediment did not result in statistically significant effects on growth,
survival, or number of mentum (mouthpart) deformities. In a separate study in C. riparius in Pb-spiked
sediment ranging from 18.1 to 456.9 mg Pb/kg dry weight, no significant differences were observed in the
frequency of mouthpart deformities (Arambourou et al„ 2012). In freshwater snail Bellamy a aeruginosa
exposed for 28 days to Pb-spiked sediment, CAT activity and metallothionein were significantly induced
at the lowest concentration tested (29.7 mg Pb/kg dry weight) (Liu et al„ 2019b). In the bivalve Hyridella
australis also exposed 28-days to Pb-spiked sediments (205 ± 9 and 419 ± 16 mg Pb/kg dry mass), the
body burden of accumulated Pb was low (2.2 ± 0.2 mg Pb/kg dry mass and 4.2 ±0.1 mg Pb/kg dry mass,
respectively); however, total antioxidant capacity significantly decreased, while ROS and MDA increased
with Pb exposure compared with controls (Marasinghc Wadigc et al„ 2014).
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As reported in the 2013 Pb ISA, inhibition of ALAD enzyme activity, an important rate-limiting
enzyme needed for heme production, is a recognized biomarker of Pb exposure in some freshwater
invertebrate species that have hemoglobin. Previous studies have indicated considerable species
differences in ALAD activity in response to Pb. For example, the concentration at which 50% ALAD
inhibition was measured in the freshwater gastropod Biomphalarict glabrcita (23 to 29 |ig Pb/L) was much
lower than that in the freshwater oligochaete L. variegatus (703 |ig Pb/L), based on nominal exposure
data (Aiscmbcrg et al., 2005). No recent studies quantifying ALAD activity in freshwater invertebrates at
environmentally relevant concentrations of Pb were identified for inclusion in this ISA. Furthermore, no
significant ALAD activity was detected at baseline metabolic conditions in hemolymph or tissue of the
freshwater unionoid mussel E. complcmata, suggesting this is not a viable biomarker for the species
(Mosheretal., 2012a).
11.3.4.3.2 Organism-Level Response
Organism-level endpoints include effects on behavior linked to Pb neurotoxicity. In the 2013 Pb
ISA, the body of evidence was sufficient to conclude there is a likely to be causal relationship between Pb
exposure and neurobehavioral effects in freshwater invertebrates (U.S. EPA, 2013) (see Table 11-4 of this
appendix). In limited studies available on worms and snails, there is evidence that Pb may affect the
ability to escape or avoid predation. For example, in the tubificid worm T. tubifex, the 96-hour EC50 for
immobilization was 42 |ig Pb/L based on nominal exposure (Khangarot, 1991). Some organisms exhibit
behavioral avoidance while others do not seem to detect the presence of Pb (U.S. EPA, 2006). Additional
behavioral endpoints reported in the Great Lakes Environmental Center draft Ambient Aquatic Life Water
Quality Criteria for Lead document U.S. EPA (2008a) include an EC50 of 140 |ig Pb/L for feeding
inhibition in the freshwater cladoceran C. dubia. In a study published since the 2013 Pb ISA, adult
amphipods, G. fossamm exposed to Pb for 5 days at a concentration at which survival was unaffected
(2.7 |ig Pb/L) exhibited sublethal behavioral and physiological responses. Locomotion was significantly
decreased overtime (assessed 24, 48 and 120 hours) and respiration rate was significantly lower at 120
hours compared with unexposed amphipods (Lebrun et al., 2017). In a separate study with G. fossarum,
both locomotion and respiration were significantly decreased following exposure to
2.1 (ig Pb/L for 24-hour (Lebrun and Gismondi, 2020).
Alterations in neurotransmitter regulation and release may be an underlying mechanism for the
behavioral effects of Pb. Few studies in freshwater invertebrates have reported effects on
neurotransmitters at lower Pb concentrations. In prereproductive freshwater bivalve Lamellidens
jenkinsiamis obesa exposed for 21 days to either 68 or 763 |ig Pb/L, AChE activity (assessed on days 1, 7,
15 and 21 of the experiment) was significantly inhibited at each timepoint compared with control
(Brahma and Gupta, 2020). Several locomotor behaviors (movement in the form of gliding, foot-siphon
extension) were significantly reduced or ceased completely in the Pb-exposed individuals compared with
the control during a separate 5-day exposure to either 69 or 776 |ig Pb/L. In the same study, reproductive-
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age individuals of another bivalve species, P. cornigata, were exposed to either 26 or 302 |ig Pb/L for
21 days. AChE activity was significantly induced at 26 |ig Pb/L and significantly inhibited compared with
control at 302 |ig Pb/L at all timepoints. Behavioral response in the form of impaired movement with Pb
exposure (25 and 304 |ig Pb/L) was also observed in this species. In 28-day chronic exposure of
freshwater snail B. aeruginosa to Pb-spiked sediment, the activity of the neurotransmitter AChE was
significantly induced starting at day 7 in the lowest concentration (29.7 mg Pb/kg dry weight) (Liu et al.,
2019b).
In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a causal relationship
between Pb exposure and growth in freshwater invertebrates (U.S. EPA, 2013) (see Table 11-4 of this
appendix). The growth of freshwater snail L. stagnalis was identified as one of the most sensitive
organisms and endpoints for Pb toxicity. At the time of the 2013 Pb ISA, the hypersensitivity of this
species to Pb was hypothesized to be from Pb inhibition of Ca2+ uptake. Subsequent experiments by Brix
et al. (2012) observed that effects on growth occur prior to effects on net Ca2+ flux, inhibition of carbonic
anhydrase activity in the snail mantle also showed no effect with Pb; therefore, the mechanism of Pb in
these highly sensitive organisms remains elusive. Additional studies reported in Section 11.3.5, Exposure
and Response of Freshwater Species, support Pb effects on the growth of L. stagnalis in the low |ig Pb/L
range (Cremazy et al., 2018; Brix et al„ 2012).
Exposure-response studies discussed in Section 11.3.5 also add to the existing body of evidence
in the 2013 Pb ISA for a causal relationship between Pb exposure and reproductive effects as well as
survival in freshwater invertebrates. In summary, studies in freshwater invertebrates for suborganism-
level and organism-level endpoints are confirmatory with findings in the 2013 Pb ISA, with evidence in
additional species for some effects.
11.3.4.4 Effects on Freshwater Vertebrates
The 1977 Pb AQCD reported Pb effects in both fish and waterfowl. The available Pb studies on
waterfowl investigated exposure to Pb via accidental poisoning or ingestion of Pb shot (U.S. EPA, 1977).
Studies on aquatic vertebrates reviewed in the 1986 Pb AQCD were limited to hematological,
neurological, and developmental responses in fish (U.S. EPA, 1986). In the 2006 Pb AQCD, effects on
freshwater vertebrates included consideration of the role of water quality parameters on toxicity to fish, as
well as limited information on the sensitivity of turtles and aquatic stages of frogs to Pb (U.S. EPA,
2006). In the 2013 Pb ISA, the body of evidence was sufficient to conclude there is a causal relationship
between Pb exposure and hematological effects, reproduction and survival in freshwater vertebrates
(based primarily on evidence from fish) (U.S. EPA, 2013) (see Table 11-4 of this appendix). There were
also likely to be causal relationships concluded between Pb exposure and physiological stress and
neurobehavioral effects. Newly available studies on the effects of Pb in fish and other freshwater
vertebrates are summarized below.
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11.3.4.4.1 Fish
11.3.4.4.2 Suborganism-Level Response
A large body of evidence supports sublethal biomarker perturbations with Pb exposure in
freshwater vertebrates; however, few studies were identified for this ISA that reported physiological
response at more environmentally relevant concentrations of Pb (< 10 |ig Pb/L; Section 11.1.1) or
concurrently assessed response at organism-level endpoints (i.e., from the cellular and subcellular level to
effects on growth, reproduction or survival). Various biomarkers of oxidative stress assessed in carp
(Cctrctssius auratus gibelio) after 96 hours and 21 days were significantly altered at analytically verified
concentration of 10 and 30 |ig Pb/L (Khan et al.. 2015). For the acute exposure, CAT activities (liver and
kidney) were significantly reduced, and SOD was significantly upregulated in brain, kidney, and muscle
tissue. GPx activity in the liver and gill increased significantly, while activity in the muscle and kidney
was significantly reduced. Biomarker response in chronic exposure showed significant reduction in CAT
(liver, gill, muscle) at 10 and 30 |ig Pb/L, whereas CAT was upregulated in the kidney. There was a
decline in GPx (liver and gill) as well as in SOD (liver, kidney, muscle), while the brain showed an
increase. Acetylcholine, a biomarker of neurotoxic stress, was significantly inhibited following chronic
exposure to 30 |ig Pb/L. Clemow and Wilkie (2015) observed no significant effect on respiratory stress,
mean cell hemoglobin concentration, plasma Ca2+ or Na2+ ion concentration or plasma protein in juvenile
rainbow trout (O. mykiss) over a 5-day exposure to 5.4 |ig Pb/L (26.1 nmol/L). In fingerling rainbow
trout, used in the same study for unidirectional Na+ flux measurement, there was an initial Na+ loss after
48 hours of exposure that recovered by 72 hours with exposure to 8.3 |ig Pb/L (40.2 nmol/L).
Hematological effects of Pb on fish reported in the 2013 Pb ISA and AQCDs include a decrease
in red blood cells and inhibition of ALAD with elevated Pb exposure under various test conditions.
Inhibition of ALAD is also reported in environmental assessments of metal-impacted habitats. For
example, as reported in the 2013 Pb ISA, lower ALAD activity has been significantly correlated with
elevated blood Pb concentrations in wild-caught fish from Pb-Zn mining areas, although there are
differences in species sensitivity (Schmitt et al.. 2007; Schmitt et al.. 2005). Few studies were identified
since the 2013 Pb ISA that quantify ALAD response in freshwater fish at concentrations considered for
this ISA (Section 11.1.1). In a field study of brown trout (Sctlmo trutta) collected from a lake in Norway
contaminated with Pb (14 (ig Pb/L) from an abandoned shooting range, ALAD activity in the trout
population was approximately 20% of that of a relatively uncontaminated reference lake (0.76 (ig Pb/L)
(Mariussen et al.. 2017).
11.3.4.4.3 Organism-Level Response
In the 2013 Pb ISA, studies supporting a likely to be causal relationship between neurobehavioral
endpoints in freshwater vertebrates and Pb exposure included research from early U.S. EPA reviews of
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the metal. In the 1977 Pb AQCD, behavioral impairment of a conditioned response (avoidance of a mild
electric shock) in goldfish was observed at concentrations as low as 70 (ig Pb/L (Weir and Hine. 1970). In
the 2006 Pb AQCD, several studies were reviewed in which Pb was shown to affect predator-prey
interactions, including alteration in prey size choice and delayed prey selection in juvenile fathead
minnows following 2-week pre-exposure to 500 |ig Pb/L (Weber. 1996). Prey capture ability was
decreased in 10-day old fathead minnow larvae born from adult fish exposed to 120 |ig Pb/L for 300 days,
then subsequently tested in a 21-day breeding assay (Mager et al.. 2010).
Since the 2013 Pb ISA, there have been additional studies on neurobehavioral response in
freshwater vertebrates, particularly in zebrafish D. rerio. As a widely used model organism in
environmental toxicology, the zebrafish genome shares a high degree of homology with the human
genome (Dai et al.. 2014; Howe et al.. 2013). Endpoints assessed in some zebrafish assays, such as
decreased locomotor activity and altered social interactions used as surrogates for autistic behaviors in
humans, can affect organism fitness in natural environments. Furthermore, some of these studies link
changes in gene expression, neurotransmitter levels or other molecular and cellular responses to the
observed behavioral outcomes. Experiments conducted in the low |ig/L range are particularly
representative of environmental concentrations (Table 11-1); therefore, zebrafish behavioral assay studies
conducted at low concentrations of Pb are reviewed below.
In zebrafish embryos exposed to 5.0, 9.7 or 19.2 |ig Pb/L there were no significant effects on
dorsal axon length up to 144 hours postfertilization (hpf); however, there was a significant reduction in
swimming speed at the highest Pb concentration tested (Zhu et al.. 2016). Alterations in the
neurotransmitter gamma-aminobutyric acid (GABA) were observed during development of zebrafish
embryos exposed to Pb (nominally to 10, 50 and 100 |ig Pb/L up to 72 hpf and then Pb uptake was
subsequently quantified in embryos) (Wirbiskv et al.. 2014). The levels of this neurotransmitter varied
with the dose of Pb and developmental stage, with all three treatments resulting in a significant decrease
in GABA by the end of embryogenesis (72 hpf). Newly hatched larval zebrafish exposed to Pb since
2.5 hpf exhibited neuromuscular responses (increased muscular twitching) at concentrations of 49.6 and
100.7 |ig Pb/L at 72 hpf. No twitches were observed at lower concentrations or in the control group
(Kataba et al.. 2022). In another study, locomotor and social behavior responses were assessed in
zebrafish larvae exposed to 4.5, 9.6 or 18.6 |ig Pb/L at 6 days postfertilization (dpf) during a dark and
light photoperiod (Zhao et al.. 2020). During the dark period, swimming activity was significantly
decreased at 18.6 |ig/L. and at both 9.6 and 18.6 |ig/L. there was a decrease in clockwise turning; social
contact time was significantly higher in the light period at the highest Pb concentration. Downregulation
of genes involved in brain neutrophic factor signaling was observed in the Pb-exposed larvae, suggesting
an underlying mechanism for the observed responses. Hyperactivity (increased distance covered and
speed) during the light period was observed in larval zebrafish exposed to Pb (3.2, 93 or 252.6 |ig Pb/L)
for 30 minutes in alternating light and dark intervals of 10 minutes (Kataba et al.. 2020).
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Studies in zebrafish have also considered the neurobehavioral effects of Pb at multiple lifestages.
Wang et al. (2018b) assessed swimming behavior in larval (15 dpf)) and juvenile (30 dpf) zebrafish that
were continually exposed to Pb (analytically verified concentration of 10 (ig Pb/L or 100 |ig Pb/L) from
maternal exposure through egg fertilization and subsequent larval development. Larval responses to Pb
exposure included decreases in measures of locomotion such as angular velocity, turn angle and inter-fish
distance, a measure of social behavior. Juvenile zebrafish exhibited similar behavioral responses to Pb;
however, the inter-fish distance increased, and there were increases in the percentage of fish moving up to
the top of the tank. The expression of key genes linked to behaviors, Ca channels and the metabolism of
environmental contaminants were altered with Pb exposure.
There is some evidence for parental transfer and transgenerational effects on fish learning and
avoidance behavior following Pb exposure. Zebrafish larvae (15 dpf) hatched from adult females
previously exposed to 19.5 |ig Pb/L were used as a model to test autism-like behaviors (Wang et al..
2016a). Behaviors assessed included measures of locomotion, and repetitive, social and anxiety
behaviors. Analysis of larval swimming activity recorded on video indicated significant increases in
distance moved and swimming velocity compared with control larvae. No significant differences were
observed in inter-fish distance, angular velocity or turn angle. Additionally, changes in the expression of
several genes associated with autism-like behaviors were detected in the larvae hatched from the Pb-
exposed fish.
In the 2013 Pb ISA, evidence was inadequate to establish a causal relationship between Pb
exposure and growth effects in freshwater vertebrates. Since the 2013 Pb ISA, a few additional studies in
fish have assessed the effects on growth following dietary or aqueous exposure to Pb. In chronic dietary
exposure (24 months) to 8-49 mg Pb/kg in food pellets, there were no significant differences in fish body
weights or the survival of Prussian carp C. gibelio females (Luszczek-Trojnar et al.. 2013). In another
study with adult female carp C. carpio exposed to Pb via diet (68.4 mg Pb/kg dry weight in food pellets),
there were no significant differences in mean body weights at the end of the study (three exposure
seasons), although Pb-exposed fish weighed significantly less than control fish after the first exposure
season (Luszczek-Trojnar et al.. 2016). This is consistent with dietary studies reviewed in the 2013 Pb
ISA (Alves and Wood. 2006). In aqueous exposure studies, zebrafish embryos exposed to Pb
(19.3 |ig Pb/L) to 6 dpf (144 hpf) showed no significant differences in hatching success, body length or
body weight compared with the control (Chen et al.. 2016b). Similarly, exposure of zebrafish embryos to
Pb (5.0, 9.7, 19.2 |ig Pb/L) up to 144 hpf did not affect growth rate or survival (Zhu et al.. 2016). No
differences in head length, head width or total body length were observed in 72 hpf embryos exposed
nominally to one of three concentrations of Pb (10, 50 or 100 |ig Pb/L and then Pb uptake was
subsequently measured in the embrvos)(Wirbiskv et al.. 2014).
For the effects of Pb on reproduction and development in freshwater vertebrates, the weight of
evidence for the causal relationship in the 2013 Pb ISA was primarily from studies with fish. Pb AQCDs
have reported developmental effects in fish, specifically spinal deformities in brook trout (Salvelinns
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fontinalis) exposed to 1 19 |ig Pb/L for three generations (U.S. EPA, 1977), as well as in rainbow trout
exposed to concentrations as low as 120|igPb/L (U.S. EPA, 1986). In the 2006PbAQCD (U.S. EPA,
2006), decreased spermatocyte development in rainbow trout was reported at 10 (ig Pb/L, and testicular
damage occurred in fathead minnow at 500 |ig Pb/L. In the 2013 Pb ISA, a 300-day chronic toxicity study
was conducted by Mager et al. (2010) in fathead minnows treated with both 31 and 112 |Lig Pb/L with
HCO3 and with 130 |ig Pb/L with DOC. The total reproductive output was decreased, and average egg
mass production increased as compared with egg mass size in controls and in low HCO3 and DOC
treatments with Pb. Other supporting evidence for the causal determination in the 2013 Pb ISA for
reproductive effects in aquatic vertebrates included alteration of steroid profiles and additional
reproductive parameters, although most of the available studies were conducted using nominal
concentrations of Pb. Additionally, a study in frogs in the 2006 AQCD showed Pb delayed
metamorphosis, decreased larval size and caused skeletal malformations at nominal concentration of
100 |ig Pb/L; however, tissue concentrations quantified in frogs following exposure fell within the range
of tissue concentrations in wild-caught tadpoles (Chen et al., 2006).
Several new early lifestage fish studies add to the existing evidence for Pb effects on endocrine
and developmental endpoints. In a study that quantified Pb in the exposure water, hatching success rates
in zebrafish embryos were reduced at 4.5, 9.6 and 18.6 |ig Pb/L. At 72 hpf, the hatching success rates in
all three concentrations were significantly decreased compared with the control, indicating that Pb caused
a hatching delay, which was also observed at the end of the experiment at 96 hpf (Zhao et al„ 2020). . In
another zebrafish study, endocrine disruption in larvae was assessed by quantifying changes in thyroid
hormone following exposure to Pb (analytically verified concentration of 2, 5, 10, 15, 20, 30 |ig Pb/L) in
embryos from 2 hpf to 144 hpf (Zhu et al., 2014). Triiodothyronine (T3) and thyroxine (T4) levels were
significantly reduced at 30 |ig Pb/L. Pb did not significantly affect the percentage of hatched larvae;
however, Pb exposure significantly increased malformations and reduced survival at 30 (ig Pb/L
compared with the control. In comparison to these studies showing reproductive and endocrine responses
in fish early lifestages, no endocrine disruption was observed in adult male common carp (C. carpio) at 7,
14 or 21 days of Pb exposure, even at the lowest analytically verified concentration (120 |ig Pb/L)
(Korkmaz et al., 2022).
Reproductive and endocrine effects of exposure to Pb via diet were assessed in dietary exposure
with female Prussian carp C. gibelio. At 12 months, there was a significant increase in luteinizing
hormone (LH) secretion after hormonal stimulation at the two highest analytically verified concentrations
(24 and 49 mg Pb/kg), whereas (8 mg Pb/kg) spontaneous LH secretion significantly decreased at the
lowest dose tested (Luszczek-Trojnar et al., 2014). At 24 months, differences in LH secretion between
treatment groups were not significant. There were also differences in oocyte size and maturation. At
12 months, oocytes in the 8 mg Pb/kg treatment group were significantly larger than those in the control
and other treatment groups. After 24 months, oocyte maturity and oocyte diameter were not significantly
different between the control and Pb-treated fish.
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11.3.4.4.4 Birds
A new study in mallards (A. platyrhynchos) expands existing information on Pb effects in birds
frequenting aquatic habitats contaminated with Pb and other metals. Prior AQCDs and the 2013 Pb ISA
include evidence for changes in ALAD activity and other oxidative stress biomarkers. A positive
relationship between the lipid peroxidation index and blood Pb in female mallards sampled in
northeastern Spain adds to this evidence. Lysozyme levels were negatively correlated with blood Pb
concentrations (Vallvcrdu-Coll et al.. 2016). Additionally, in male mallards, there were significant
relationships between blood Pb and beak and leg hue. In mallards, male leg and beak color typically
ranges from orange-red to yellow-orange and from yellow-orange to green, with redder beaks and
yellower legs typically being more attractive to females. In this study, the leg redness of males had a
significant negative relationship with blood Pb levels, as did beak yellowness. This indicates that male
mallards with higher blood Pb levels are likely to be less attractive to females, and therefore could
potentially have lower reproductive success. Another study from the same author investigated how blood
Pb levels in mallard chicks can affect multiple suborganismal and organismal-level effects (Vallvcrdu-
Coll et al.. 2015). This study on the same population of mallards in northeastern Spain found that
ducklings with blood Pb levels above 180 ng/mL showed reduced body mass and died during the first
week posthatching. Additionally, cellular immune function at day 15 in ducklings was negatively
correlated with Pb levels in blood on the same day.
11.3.4.4.5 Amphibians
Since the 2013 Pb ISA, new laboratory studies on the effects of Pb exposure on freshwater
amphibians have focused on tadpole growth, development, and survival. Asiatic toads reared in water
with different concentrations of Pb (0, 9.85, 48.73, 97.69, 497.34, and 998.27 |ig Pb/L analytically
verified concentrations for 0, 10, 50, 100, 500, 1,000 |ig Pb/L) showed a significant increase in total
length and body mass at 50 (ig Pb/L and a significant decrease in snout-to-vent length at 1000 |ig Pb/L on
day 10 compared with controls. However, farther along in development at day 20, there was a significant
decrease in snout-to-vent length at 100 and 500 |ig Pb/L compared with controls (Yang et al.. 2019).
Huang et al. (2014) examined the effect of Pb on these endpoints in dark-spotted frogs
(Pelophylax nigromaculata). Tadpoles were reared in different concentrations of Pb 38.2, 79.3, 158.4,
318.7, 638.1, 1278.9 (ig Pb/L (analytically verified concentrations for 40, 80, 160, 320, 640,
1280 |ig Pb/L) from heartbeat to complete tail reabsorption. The threshold concentrations for effects on
body mass, snout-vent length, forelimb length, and hindlimb length were 160, 160, 160, and 320 |ig Pb/L,
with total malformation rate increasing linearly with Pb concentration. Metamorphosis time was
significantly affected by Pb concentration and exhibited a linear increase with increasing Pb concentration
(0 ng Pb/L = 76.4 ± 0.5 days, 160 ^g Pb/L = 90.8 ± 0.5 days, 1280 ^g Pb/L = 118.4 ± 0.5 days). Pb
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concentration also significantly affected the survival rate, which decreased with increasing Pb
concentration (0 |ig Pb/L = 98.3 ± 1.7%, 160 |ig Pb/L = 93.3 ± 1.7%, 1280 |ig Pb/L = 80.0 ± 0.3%).
Other than the studies in fish described above and in the following section on exposure-response,
there is limited new information regarding Pb toxicity in freshwater vertebrates. For fish, studies are
largely confirmatory with studies in the 2013 Pb ISA. Additional research with zebrafish augment
existing understanding of Pb effects on neurobehavior and reproductive endpoints.
11.3.5 Exposure and Response of Freshwater Species
Evidence regarding exposure-response relationships and potential thresholds for Pb effects on
aquatic populations can provide tools for quantitative analyses of risks in freshwater ecosystems
(Section 11.1.7.3). Exposure-response data for the reproduction, growth, and survival of freshwater biota
(including microalgae, invertebrate, amphibian, and fish species) were summarized in Table 6-5 of the
2013 Pb ISA (U.S. EPA. 2013). Additionally, the Annex of the 2006 Pb AQCD (U.S. EPA. 2006)
summarized data on exposure-response functions for invertebrates (Table AX7 2.4.1) and fish
(Table AX7 2.4.2) available at the time. For Pb exposure-response, there is significant new research
reporting results from bioassays of freshwater algae, invertebrates and fish based on measured rather than
nominal concentration of Pb. In some cases, effects were observed in sensitive species at concentrations
comparable to or lower than those reported in the 2013 Pb ISA (Table 11-5) or earlier U.S. EPA reviews
of Pb. Some of the studies report LCio and LC20 toxicity values and/or calculate the free-ion
concentration.
In the 2006 AQCD and 2013 Pb ISA, available exposure-response data for freshwater plants and
algae did not indicate any effects on growth or survival at environmentally relevant concentrations. In the
2006 AQCD, EC50 values for growth inhibition in various freshwater algal and aquatic plant species were
between approximately 1,000 and >100,000 |ig/L and were mostly based on nominal concentration data
(U.S. EPA, 2006). An important advancement since the 2013 Pb ISA is the availability of bioassay data
for algal growth rate in several freshwater species based on measured Pb concentration instead of nominal
concentration, which strengthens confidence in the findings for the concentrations assessed (De
Schamphelaere et al., 2014). In chronic 72-hour bioassays in standard test media to assess the growth rate
in three commonly tested algal species (P. subcapitata, C. kesslerii, C. reinhardtii), P. subcapitata was
the most sensitive, with EC50 = 83.9 |ig Pb/L, EC20 = 45.7 |ig Pb/L and EC10 = 32.0 |ig Pb/L based on
filtered Pb concentrations (De Schamphelaere et al„ 2014). Furthermore, in subsequent tests with P.
subcapitata at varying pH, the 72h EC50 decreased from 72.0 |ig filtered Pb/L at pH 6.0 to 20.5 |ig filtered
Pb/L at pH 7.6. Inhibitory concentration (IC) values calculated using a specific growth rate at 72 hours
with a linear interpolation method for Raphidocelis subcapitata (formerly P. subcapitata) were
IC10 = 0.15 (j,M, (31 ng Pb/L), IC25 = 0.39 (81 ^g Pb/L) and IC50 = 0.78 (161 ^g Pb/L) (Alho et
al.. 2019).
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In addition to freshwater algae, there is new toxicity information based on measured Pb
concentration for freshwater plants. The toxicity of Pb to duckweed Lemnct minor expressed as percent
net root elongation was assessed in chronic bioassays of seven U.S. surface waters with different water
chemistries (Antunes and Kreager, 2014). The 20% IC in 7-day static renewal tests with the waters ranged
from 306 nM to >6920 nM (63 |ig Pb/L to >1,433 |ig Pb/L) expressed as total dissolved Pb indicating that
Pb speciation, solubility, subsequent bioavailability, and toxicity varied under the range of water
hardness, pH, and DOC in the tested waters.
For freshwater invertebrates, effects in sensitive species of amphipods, gastropods, cladocerans
and mussels were reported at low |ig Pb/L concentrations in exposure-response studies reviewed in the
1986 AQCD, the 2006 AQCD and the 2013 Pb ISA. Additional toxicity data for these taxonomic groups
discussed below support and expand upon what was known in the previous Pb assessment in terms of the
relative sensitivity of these freshwater biota to Pb.
Toxicity testing with amphipods reported in the 2006 AQCD and 2013 Pb ISA indicate a
response to Pb at <10 |ig Pb/L under some water conditions. At higher pH and water hardness, these
organisms are less sensitive to Pb (U.S. EPA, 2006). For example, a 7-day LC50 of 1 |ig Pb/L was
observed in soft water with the amphipod H. azteca (Borgmann et al., 2005). In this same species, the 96-
hour LC50 for Pb at pH 5 was 10 |ig Pb/L (Mackie, 1989). In 42-day chronic exposures of H. azteca
exposed to Pb via water and diet, the LC50 was 16 |ig Pb/L (Besser et al., 2005). In a chronic 42-day
bioassay with H. azteca, published after the 2013 Pb ISA, survival was similar to that observed by Besser
et al. (2005) under two different experimental diets conducted concurrently (LC20 =15 |LLg Pb/L and
LC20 =13 |Lig Pb/L) and support the findings of effects in amphipods in the low |ig/L range (Besser et al.,
2016).
Some species of freshwater gastropods have exhibited sensitivity to Pb at <20 |ig Pb/L. In the
1986 AQCD, Borgmann et al. (1978) found increased mortality at Pb concentration as low as 19 |ig Pb/L
in the freshwater snail Lymnaea palutris exposed from hatching to reproductive maturity (approximately
120 days). To follow-up on the set of studies reviewed in the 2013 Pb ISA (Grosell and Brix, 2009;
Grosell et al., 2006b) that identified the freshwater snail L. stagnalis as highly sensitive to Pb
(EC20 = <4 |ig Pb/L in 30-day exposure experiments) several additional chronic studies have since been
undertaken with this species. In growth bioassays conducted in a variety of natural waters across the
United States with different water chemistries 14-day EC20 and EC50 values ranging from 1.5 to 49.5 and
from 3.6 to 244.6 |ig Pb/L, respectively, were reported forZ. stagnalis (Esbaugh et al„ 2012). Munley et
al. (2013) conducted full lifecycle bioassays with a duration of 56 days to assess the effects on survival,
growth, reproduction, and development in L. stagnalis and determine if there was any recovery from
growth inhibition effects reported in the 30-day exposures. Survival was significantly decreased at the
highest concentration of Pb tested (8.4 |ig Pb/L) after 21-days of exposure until the end of the experiment,
for a final NOEC = 2.7 |ig Pb/L and LOEC = 8.4 |ig Pb/L. Consistent with the earlier 30-day exposures,
growth was significantly decreased at day 28, even at the lowest tested concentration (1.0 (ig Pb/L), for
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NOEC < 1.0 |ng Pb/L and LOEC = 1.0 jag Pb/L. By day 56, growth remained significantly lower than that
of the controls in the 2.7 and 8.4 |ig Pb/L concentration; however, snails exposed to 1.0 |ig Pb/L
surpassed the growth rates of the unexposed snails. Inhibition of the specific growth rate at the
2.7 |ig Pb/L exposure was observed during the last week of the experiment. Conducting a 56-day lifecycle
bioassay with L. stagnctlis enabled assessment of reproductive and developmental endpoints (Munlcv et
al.. 2013). The reproductive phase started at day 32 and continued till the end of the study. For the
number of egg masses and time until first egg mass, the NOEC <1.0 jag Pb/L and LOEC = 1.0 |ig Pb/L.
No effects of Pb on the number of embryos per egg mass were observed at any concentration tested.
Individuals exposed to the highest concentration (8.4 |ig Pb/L) did not reproduce during the lifecycle test.
Egg capsule and embryo diameters after 7 days of development were significantly reduced at 2.7 |ig Pb/L
(the highest concentration in which snails reproduced in the study). Although growth exhibited some
recovery in L. stagnctlis in the longer 56-day lifecycle tests, growth effects observed at 28 days were
predictive of the reproductive effects observed in the longer exposure (Munlev et al.. 2013). Additional
growth studies conducted by Brixet al. (2012) reported an EC20 (biomass) at 8 days of exposure of 3.2 |ig
1_1 Pb and 3.5 |ig 1 1 Pb after 16 days of exposure. Under similar experimental conditions. Cremazv et al.
(2018) reported a 14-day EC10 of 4 |ig Pb/L, an EC20 of 7.67 |ig Pb/L and an EC50 of 23.4 |ig Pb/L for
juvenile growth from compiled results of multiple toxicity tests. The corresponding chronic growth effect
concentrations based on free-ion activity were EC10 = 0.157 |ig Pb/L, EC20 = 0.320 |ig Pb/L and
ECso = 1.08 |ig Pb/L.
New acute data for cladocerans include a 48-hour EC50 = 280 |ig Pb/L for immobilization in D.
magna (Okamoto et al.. 2015). Among the studies reviewed in the 2013 Pb ISA was a series of 48-hour
acute toxicity tests using a variety of natural waters across North America. The cladoceran C. dubia. LC50
values in that study ranged from 29 to 180 |ig Pb/L, and DOC was well correlated with protection against
the toxicity of Pb (Esbaugh et al.. 2011). In this same species, increasing DOC led to an increase in the
mean EC50 for reproduction, ranging from approximately 25 |ig Pb/L to >500 |ig Pb/L in 7-day chronic
toxicity bioassays (Mager et al.. 201 la). In a study published after the 2013 Pb ISA in this same species, a
series of 7-day reproductive toxicity tests to assess the effects of metal mixtures reported an EC50 range of
111 to 302 |ng Pb/L in the Pb-only treatments (Nvs et al.. 2016a). In another study with C. dubia, the EC50
for reproduction ranged from 99.8 |ig Pb/L at pH 6.4 to 320 |ig Pb/L, at pH 8.2, and 81.2 (ig Pb/L at
0.25 mM Cato 130 (ig Pb/L at 1.75 mM Ca (Nvs et al.. 2014). In comparison, in a series of chronic Pb
toxicity tests conducted in a variety of natural waters across the United States with different water
chemistries which expanded upon the findings of Esbaugh et al. (2011). 7-day EC20 for reproduction in C.
dubia ranged from 12.1 to 223.3 |ig Pb/L, and 7-day-EC5o ranged from 20.1 to 573.4 |ig Pb/L (Esbaugh et
al.. 2012).
Using the same set of waters from across the United States, reproduction (as population growth)
was also assessed in rotifer/', rapida over a 4-day exposure period (Esbaugh et al.. 2012). Chronic EC20
and EC50 in this species based on dissolved Pb concentration ranged from 3.2 to 103.3 and 10.6 to
154.9 |ig Pb/L, respectively. The variability in toxic response to Pb was linked to water chemistry; DOC
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had a protective effect for C. dubia and snail L. stagnalis, while rotifer response was most closely
associated with Ca and pH, not DOC. In comparison, another species of rotifer, B. calyciflorus, was less
sensitive to Pb; 4-day chronic reproductive toxicity EC20 ranged from 75 |ig Pb/L to 336 |ig Pb/L and
EC50 ranged from 138 to 634 |ig Pb/L in natural waters of varying chemistry (Nvs et al.. 2016b).
In response to a lack of chronic toxicity data in freshwater isopods based on measured
concentrations Van Ginneken et al. (2017) conducted a series of exposure-response studies with trace
metals including Pb in adult A aquations. The authors determined LC10, LC20 and LC50 effect values for
this species (14-day LC10 = 49.7 |ig Pb/L, LC20 = 130 |ig Pb/L, LC50 = 677 |ig Pb/L) and also calculated
lethal concentrations based on free-ion activity using the Windermere Humic Aqueous Model
(LC10 = 0.04 |ig/L. LC20 = 0.31 (ig/L and LC50 = 9.13 (ig/L). In a separate study w ith A. aquations, the 10-
day LC50 was 443 |ig Pb/L (Van Ginneken et al.. 2015). In another crustacean, juvenile prawns (M
nipponense), no statistically significant effects on mortality were reported at 12 or 25 |ig Pb/L
concentration in chronic 60-day exposure trials; however, reductions in weight gain and specific growth
rate were observed in the prawns exposed to 25 |ig Pb/L (Ding et al.. 2019).
In freshwater mussels, sensitivity to Pb has been demonstrated to vary with lifestage. In a study
from the 2013 Pb ISA, newly transformed juvenile freshwater mussels (Lampsilis siliqiioidea) were more
sensitive than older juveniles in acute exposures. A chronic value (geometric mean of NOEC and the
LOEC) of 10 |ig Pb/L was reported in 28-day exposures of 2-month-old juveniles (Wang et al.. 2010).
The lowest median effect concentration for glochidia (larvae) of L. siliqiioidea at 24 and 48 hours was
>299 |ig/L. A more recent study in glochidia of six different freshwater mussel species found in
southeastern Australia (Hyridella australis, Hyridella depressa, Velesunio ambigmis, Alathvria profuga,
Ciiciimeriinio novaehollandiae, Hyridella drapeta) indicated these species were more sensitive in acute
tests than glochidia of L. siliqiioidea (native to the United States). The 24-hour EC50 values for valve
closure ranged from 176 to 274 |ig Pb/L (Markich. 2017). Following 72-hour Pb exposure in the same
species, the EC50 values ranged from 65 to 110 (ig Pb/L. Calculated no-effect concentrations (NECs) at
72 hours ranged from 11 to 21 |ig Pb/L.
Other recent tests with freshwater invertebrates have illustrated the range in the sensitivity of
North American species to Pb. In a battery of acute toxicity tests using resident invertebrates collected
from the South Fork Coeur d'Alene River watershed, Idaho, U.S. and tested in the river water, the lowest
EC50 concentration for Pb (96-hour EC50 = 253 |ig Pb/L) was obtained with the stonefly Sweltsa sp.,
however, in other tests with Sweltsa sp., mortalities occurred at Pb concentrations up to three times
greater, indicating a high degree of variability in repeated tests with the same species (Mebane et al..
2012). Additional invertebrates were tested in waters from the South Fork Coeur d'Alene River
watershed, Idaho, U.S., and their lowest corresponding 96-hour EC50 values (some invertebrate species
were tested multiple times) were: four mayfly species (Baetis tricaudatus [96-hour LC50 = 322 to
<1,250 |ig Pb/L tested at varying water hardness], Rhithrogena sp. [96-hour LC50 = >166 |ig Pb/L],
Drunella sp. [96-hour LC50 = >267 |ig Pb/L], Epeorus sp. [96-hour LC50 = >346 |ig Pb/L] and
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Leptophlebiidae [96-hour LC50 = >346 |ig Pb/L]), a caddisfly (Arctopsyche sp. 96-hour
LC50 = >1,255 |ig Pb/L), a Simuliidae black fly (96-hour LC50 = 415 |ig Pb/L), Chironomidae midge (96-
hour LC50 = 1,955 (ig Pb/L), a Tipula sp. Crane fly (96-hour LC50 = >1,035 |ig Pb/L), a Dytiscidae beetle
(96-hour LC50 = >1,035 |ig Pb/L) and two snail species (Physa sp. [96-hour LC50O = 1,159 |ig Pb/L] and
Gyraulus sp [96-hour LC50 = 380->l,035 |ig Pb/L] tested at varying water hardness).
Since the 2013 Pb ISA, additional exposure-response information has been obtained from
sediment bioassays for freshwater invertebrates. In 21-day whole sediment chronic toxicity bioassays, no
negative effect was noted for larvae of the North American mayfly species, Hexagenia limbata, exposed
up to 2,903 mg Pb/kg sediment (highest concentration tested); for survival, the porewater
LOEC = >130 |ig/L and overlying water LOEC = >53.6 |ig/L (Nguyen et al., 2012). In the same study, for
a European species Ephoron virgo, 21-day EC50 and LOEC of 2,201 and 2,071 mg Pb/kg were found,
respectively, with a porewater LOEC = 105 |ig Pb/L and overlying water LOEC = 19 |ig Pb/L. In long-
term whole-sediment toxicity tests with three benthic organisms exposed to various concentrations of Pb;
L. variegcttus (16 to 5,746 mg Pb/kg), G. piilex (21 to 2,734 mg Pb/kg) and mayfly Ephoron virgo (15 to
2,972 mgPb/kg), in which Pb-spiked sediments were allowed to fully equilibrate 35 or 40 days prior to
testing and metal concentrations were monitored throughout, the survival of E. virgo (21-day
EC10 = 1,455 mg Pb/kg dry weight) and the biomass of L. variegcitus (28-day EC10 = 1,870 mg Pb/kg dry
weight) were more sensitive endpoints compared with the growth of G. pulex (35-day
EC10 = 2,541 mg Pb/kg dry weight) (Vandcgchuchtc et al., 2013).
For freshwater vertebrates, the majority of available exposure-response data are for fish. In the
studies reviewed for the 2006 Pb AQCD, freshwater fish demonstrated negative effects at concentrations
ranging from 10 to >5,400 |ig Pb/L, generally depending on exposure duration and water quality
parameters (e.g.,pH, hardness, salinity) as summarized in Table AX7 2.4.2 of the 2006 AQCD (U.S. EPA,
2006). In the 2013 Pb ISA, several acute and chronic bioassay studies with fish further elucidated the role
of water chemistry in toxicity (Esbaugh et al„ 2011; Grosell et al„ 2006b; Grosell et al„ 2006a). In a
series of 96-hour acute toxicity tests with fathead minnow (P. promelas) conducted in a variety of natural
waters across North America, LC50 values ranged from 41 to 3,598 |ig Pb/L in this species
(Esbaugh et al., 2011). Chronic assays with rainbow trout reported in the 2013 Pb ISA provided
additional exposure-response data for this species. In a 69-day test with rainbow trout, the following
chronic values were observed for survival: NOEC = 24 |ig Pb/L, maximum acceptable toxicant
concentration (MATC) = 36 |ig Pb/L, EC10 = 26 |ig Pb/L, EC20 = 34 |ig Pb/L and LC50 = 55 |ig Pb/L
(Mebane et al„ 2008). Results from a 62-day test, with fish length as the endpoint, were
NOEC = 8 ng Pb/L, MATC = 12 |ig Pb/L, EC10 = 7 jig Pb/L, EC20 = 102 |ig Pb/L and LCso = 120 |ig Pb/L
(Mebane et al„ 2008).
New evidence since the 2013 Pb ISA includes additional studies on fish species native to North
America. In 96-hour acute toxicity tests with white sturgeon (A. transmontanus), which is experiencing
population declines in the United States and Canada, two early lifestages (8 and 40 dph) were tested in lab
11-139
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water and in water from the Columbia River upstream of the Teck Trail smelter facility, British
Columbia, Canada (Vardv et al.. 2014). For 8 dph larvae, 96-hour LC50 =177 |ig/L (lab water) and 96-
hour LC50 >410 |ig/L (river water); for 40 dph, 96-hour LC50 = 528 |ig/L (lab water) and 96-hour
LC50 = 1,556 |ig/L (river water) (Vardv et al.. 2014). In 27 dph juvenile white sturgeon exposed to Pb
concentrations in water ranging from 0.03 to 60 |ig Pb/L for 28 days, there was an EC20 > 60 |ig Pb/L for
survival, length, and biomass (Wang et al.. 2014a). Considering that the early lifestages of white sturgeon
are in close contact with sediment and porewater Balistrieri et al. (2018) reported an EC20 = 0.9 nM Pb2+
(0.18 |ig Pb2+/L) developed from predictive response modeling using in situ measurements of Pb in
Columbia River sediment and porewater, free-ion concentrations from equilibrium speciation calculations
and the laboratory toxicity testing results of Wang et al. (2014a) of Pb to the early lifestages of sturgeon.
Similar dose-response curves based on free metal ion concentration were observed for effective mortality
and for reduction in biomass at Pb2+ concentrations higher than quantified in sediment porewater,
indicating young sturgeon at the sediment-water interface are unlikely to be affected by toxic
concentrations of Pb in the upper reaches of the Columbia River. Mebane et al. (2012) tested westslope
cutthroat trout (Oncorhvnchus clarkii lewisi) a native subspecies of conservation concern, in a series of
bioassays using water from various locations within the South Fork Coeur d'Alene River watershed,
Idaho. EC50 values for the effective mortality for this species ranged from 47 to 487 |ig Pb/L.
In native rainbow trout (O. mykiss), 7-week waterborne-only exposure (4, 11,21, 82, 251 and
907 |ig Pb/L) conducted as part of a larger study to assess the toxicity of different dietary pathways in
juvenile rainbow trout, survival was assessed daily, and fish were weighed weekly (Alsop et al.. 2016). At
96-h, toxicity values were LC10 = 304.3 |ig Pb/L, LC20 = 357.7 |ig Pb/L and LC50 = 487.3 |ig Pb/L. At
7 weeks, LC10 = 55.6 |ig Pb/L, LC20 = 96.9 |ig Pb/L and LC50 = 280.2 |ig Pb/L. All fish exposed at the
highest concentration did not survive, and no significant effects on growth were reported for any
concentration for the duration of the experiment. In 27 dph juvenile rainbow trout, EC20 > 128 |ig Pb/L
for survival, length and biomass following 28 days of Pb exposure (Wang et al.. 2014a). In tests with
larval trout, EC20 values were the same as observed in the juveniles. In addition to studies on native fish
species, other studies in fish support previous understanding of the role of water chemistry in Pb toxicity.
For larval zebrafish (D. rerio) acute toxicity, 96-hhour LC50 values varied with water hardness; in soft
water LC5o= 52.9 |ig Pb/L and in hard water LC50 = >590 |ig Pb/L (Alsop and Wood. 2011). |ig Pb/L
(Alsop and Wood. 2011).
As discussed in Section 11.1.7.3, the existing U.S. EPA AWQC for Pb for the protection of
aquatic life are CMC of 65 |ig Pb/L (for acute exposure) and CCC of 2.5 |ig Pb/L (for chronic exposure)
at a hardness of 100 mg/L (U.S. EPA. 1985a). Since these criteria were developed in 1984, there have
been additional acute and chronic toxicity data and improved characterization of modifying factors that
affect Pb bioavailability and toxicity. Taking these advances into consideration Deforest et al. (2017)
proposed updated acute BLM-based aquatic life criteria, ranging from 18.9 to 998 |ig Pb/L and chronic
BLM-based Pb freshwater criteria ranging from 0.37 to 41 |ig Pb/L (Table 11-5). The lowest criteria were
for water with low DOC (1.2 mg/L), pH (6.7) and hardness (4.3 mg/L as CaCCh), and the highest criteria
11-140
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were for water with high DOC (9.8 mg/L), pH (8.2) and hardness (288 mg/L as CaCCh). which
encompasses varying water quality conditions of North American surface waters and the importance of
DOC and pH as modifying factors compared with hardness. The updated data sets in Deforest et al.
(2017) incorporated toxicity information for L. stagnalis, C. dubia, H. ctztecct and P. rctpida, freshwater
invertebrates that are relatively sensitive to Pb exposure. The number of genera with acute toxicity data
for Pb increased from 10 to 32, and for chronic toxicity, from 4 to 13, which enabled the proposed chronic
criteria to be based on bioassay data rather than an acute-to-chronic ratio that was used in 1984 for
derivation of the CCC.
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Table 11-5
Studies in freshwater biota with analytically verified Pb concentrations and that report an effect on
growth, reproduction or survival comparable to, or lower than, the lowest effect concentrations
reported in previous Pb AQCDs or the 2013 Pb ISA
Species
Concentration Exposure Method
Modifying
Factors
Effects on Endpoint
Effect Concentration
Reference
(published
since the
2013 Pb ISA)
Algae/Plants
Green algae
(Pseudokirchneriella
subcapitata),
Green algae
(Chlorella kessleri)
Green algae
(Chlamydomonas
reinhardtii)
P. subcapitata
Total Pb:
<1, 19, 42, 85,
228.5,
412 Pb |jg/L
Filtered Pb:
<1, 16, 37, 77,
201, 418 Pb |jg/L
C. kesslerii
Filtered Pb:
<1, 7, 18, 39, 80,
164, 417 |jg Pb/L
C. reinhardtii
Filtered Pb: <0.8,
9.5, 19.8, 43.3,
89.4, 194, 452,
783, 1,613 |jg
Pb/L
Standard 3-d toxicity
tests conducted in
OECD standard test
medium with addition
of 4 mg/L of
Suwannee River
Fulvic Acid. Cell
densities were
measured after 24,
48 and 72 h of
exposure using a
particle counter. The
growth rates of C.
vulgaris and C.
reinhardtii were not
considered
exponential during
the third day of
exposure, so the 2-d
ECx values were
calculated for these
two species.
Additional tests were
conducted with P.
subcapitata with
varying pH and fulvic
acid
Temperature:
24°C
pH =6
Growth:
Interspecies comparison of algal
growth rate indicated that P.
subcapitata is the most sensitive
and C. kesslerii the least
sensitive. In P. subcapitata, as
pH increased from 6.0 to 7.6, the
72-hr ECso decreased from 72.0
to 20.5 |jg filtered Pb/L
P. subcapitata
2-d ECso = 89.9 pg Pb/L
2-d EC20 = 44.7 |jg Pb/L
2-d EC10 = 29.7 |jg Pb/L
3-d ECso = 83.9 pg Pb/L
3-d EC20 = 45.7 pg Pb/L
3-d EC10 = 32.0 pg Pb/L
C. kesslerii
2-d ECso = 388 pg Pb/L
2-d EC2o= 185 pg Pb/L
2-d ECio= 120 pg Pb/L
C. reinhardtii
2-d ECso = 172 pg Pb/L
2-d EC2o= 108 pg Pb/L
2-d EC10 = 82.3 pg Pb/L
De Sch-
amphelaere
etal. (2014)
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
Green algae
(Raphidocelis
subcapitata formerly
known as
Pseudokirchneriella
subcapitata)
(1.20; 2.41; 4.82
and 12.06 pM)
Nominal, stock
solution
analytically
verified
72-hr toxicity test
with triplicates,
maintained in a
temperature-
controlled room. Cell
density assessed
every 24 h
Temperature: Growth:
25 ± 2°C Pb significantly inhibited algal
growth. All treatments differed
significantly (p < 0.05) from the
control group at 72 h of
exposure. Pb completely
inhibited algal growth at
12.06 pM
72-hr IC10 = 0.15 pM,
(31 pg Pb/L)
72-hr IC25 = 0.39 pM
(81 pg Pb/L)
72-hr IC50 = 0.78 pM
(161 pg Pb/L)
Alho et al.
(2019)
Duckweed
(Lemna minor)
A range of
concentrations as
low as 10 pg Pb/L
to as high as
9,740 pg Pb/L.
Total Pb added to
each water was
varied because
waters differed in
hardness, DOC,
and pH. All
waters were
equilibrated for
24 h prior to
bioassays
A series of 7-d static
renewal tests with L.
minor were
conducted with
seven different
surface waters
collected from across
the United States
with varied
chemistries and
spiked with a
concentration series
of Pb(N03)2. Plants
were held in a growth
chamber and growth
was assessed as %
net root elongation
Temperature:
25 ± 2°C
pH:
5.4-8.3
depending on
surface water
DOC:
0.5-12.5 mg/L
depending on
surface water
Hardness:
8-266 mg/L
CaC03 depending
on surface water
Growth: The inhibition of net root 20% inhibitory
elongation varied widely
depending upon the chemistry of
the assayed waters and its
effects on Pb speciation
concentration in 7-d static
renewal tests with the
waters ranged from
306 nM (63 pg Pb/L) to
>6920 nM to
(>1,433 pg Pb/L) total
dissolved Pb
Antunes and
Kreaaer
(2014)
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Species
Concentration Exposure Method
Modifying
Factors
Effects on Endpoint
Reference
Effect Concentration (published
since the
2013 Pb ISA)
Invertebrates
Amphipod
(Hyalella azteca)
Control 5, 10, 20,
40, 80 |jg Pb/L.
Pb aqueous
concentrations
varied among diet
treatments and
overtime,
suggesting that
food inputs
modified Pb
concentration and
bioavailability
7-d-old amphipods in
flow-through water-
only exposure to Pb
as Pb-nitrate in 42-d
chronic bioassays.
Amphipods were fed
one of two
experimental diets: a
suspension of YCT
or a DT fish food diet.
Assays conducted
concurrently in test
water from the same
diluter system
Hardness
100 mg/L as
CaCC>3
pH about 8.2
Alkalinity 95 mg/L
Survival:
Survival was similar with
aqueous Pb exposure in
amphipods fed two different diets
Growth:
Biomass significantly reduced in
amphipods fed YCT, not
significantly reduced in
amphipods fed DT up to
63 |jg Pb/L
Reproduction:
Fecundity significantly reduced in
amphipods fed YCT, not
significantly reduced in
amphipods fed DT up to
63 |jg Pb/L. (Note: fecundity and
total young endpoints did not
meet test acceptability criteria for
YCT diet).
Lowest reliable toxicity Besser et al.
value for each endpoint in (2016)
|jg/L filtered Pb:
DT diet:
42-d EC2o= 13 |jg Pb/L
42-d NOEC = 5.9 pg Pb/L
42-d LOEC = 13 |jg Pb/L
YCT diet:
42-d EC2o= 15 |jg Pb/L
42-d NOEC = 6.1 |jg Pb/L
42-d LOEC = 14 |jg Pb/L
Lowest biotic ligand
model-normalized effect
concentrations:
EC20 = 8.2 |jg Pb/L (total
young for the DT test)
ECso = 6.6 |jg Pb/L
(biomass for the YCT test)
Isopod
(Asellus aquaticus)
15.1, 31.1, 74.7,
203, 443 |jg Pb/L
Various metal
mixtures and single
metals were tested in
a 10-d exposure with
individuals of equal
length
(9.43 ± 0.17 mm) in a
climate chamber.
The Pb-only
treatment was Pb as
PbCb
Temperature:
20 ± 1°C
Hardness:
117 mg L"1
CaC03
Survival: Focus of study was on
mixture toxicity. Only LC50 was
calculated for Pb-only treatment
10-d LC50 = 443 pg Pb/L Van Ginneken
et al. (2015)
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
Isopod
(Asellus aquaticus)
0.71 (control),
25.6, 110, 358
and
37,616 |jg Pb/L
(measured values
for effective
concentration)
<0.1, <0.1, 0.36,
4.67 and
18,982 |jg Pb/L
(free-ion activities
of the measured
effective
concentrations
calculated using
the Windermere
Humic Aqueous
Model with 100%
of DOC as fulvic
acids)
Chronic 14-day
exposure to
Pb(NC>3)2 with adult
A. aquaticus. Assay
water sampled on
days 0,1,4, 7 and 14,
isopods were
removed from
exposure containers
for 4 h on day 7 for
feeding
Temperature:
15 ± 1°C
pH:
7.72 ± 0.03
DOC:
5.94 ± 0.13 mg/L
Dissolved oxygen:
8.68 ± 0.03 mg/L
Survival: Severe mortality was
only observed at the highest
concentration tested after 14-d
exposure. Low mortality was
observed in the other
concentrations. During the
exposure period, LC values
declined until day 4, then
continued to slowly decrease.
The free-ion activities produced
the lowest LC values
14-d survival
LC10 = 49.7 |jg Pb/L
LC10 for
FIA = 0.04 |jg Pb/L
LC2o= 130 |jg Pb/L
LC20 for
FIA = 0.31 |jg Pb/L
LC50 = 677 |jg Pb/L
LC20 for
FIA = 9.13 |jg Pb/L
7-d survival
LC10 = 97.4 |jg Pb/L
LC20 = 602 |jg Pb/L
LC50 = 13,562 |jg Pb/L
(LC10, 20 and 50 values
were also calculated for
day 1 and day 4).
Van Ginneken
etal. (2017)
11-145
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
Cladoceran
(Ceriodaphnia
dubia)
pH 6.4, 7, 7.6
series: (nominal
concentration 80,
110, 140, 170,
220, 320 |jg Pb/L)
pH 8.2 test:
(nominal
concentration
100, 160 220 280,
340, 400 |jg Pb/L)
Ca test series:
(nominal
concentration 50,
100, 150, 220,
320, 400 |jg Pb/L)
Total and filtered
Pb in each series
quantified but not
reported for
individual assays
Reproductive effects
of Pb (PbCh)
assessed in 7-d
chronic assays.
Juveniles (<24 h old)
exposed to Pb and
varying Ca or pH in
static renewal
assays. Mortality and
number of juveniles
noted daily
PH
4 series:
6.4; 7; 7.6; 8.2
Hardness
4 series:
Ca = 0.25 mM;
1.0 mM; 1.75 mM;
2.5 mM
DOC
3.2-3.3 mg/L in
pH series
3.8-4.0 in
hardness series
Reproduction
Total reproduction (number of
juveniles per female) relative to
the mean control reproduction
varied with Ca or pH over 7-d
chronic exposure to Pb. High pH
was protective of Pb toxicity and
water hardness had less effect
on chronic toxicity than pH
7-d ECso for reproduction
ranged from 99.8 |jg Pb/L
at pH 6.4 to 320 ug Pb/L
at pH 8.2
7-d EC50 for reproduction
ranged from 81.2 |jg Pb/L
at 10 mg/L (0.25 mM) Ca
to 130 |jg Pb/L at 70 mg/L
(1.75 mM) Ca
Nvs et al.
(2014)
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Cladoceran
(Ceriodaphnia
dubia)
Rotifer
(Philodina rapida)
Snail (Lymnaea
stagnalis)
Each species was
tested in a range
of concentrations
starting at low
|jg Pb/L. Actual
concentrations
were measured
but not reported
for the individual
assays
All three species
exposed to Pb as
Pb(NC>3)2, in a range
of representative
surface waters
across North
America
C. dubia\ (<24-hr-old
neonates) 7-d
chronic reproductive
bioassays conducted
in a temperature-
controlled chamber
with a combination of
dietary and aqueous
exposure and
monitored daily for
survival and
reproduction
P. rapida\ 4-d chronic
Pb toxicity with adults
was assessed using
a population growth
rate endpoint which
conformed to
classical
concentration-
dependent
responses.
Representative
surface waters for
the bioassays had
varying pH, DOC,
and water
hardness
C dubia:
pH: 6.51-8.47
DOC: 114-1443
Temperature:
26°C
P. rapida:
pH: 7.23-8.44
DOC: 79-1405
Temperature:
26°C
L. stagnalis:
pH: 5.79-8.61
DOC: 36-1314
Temperature:
26°C
Reproduction:
Highest reproductive toxicity in C.
dubia was observed in soft water,
most protective water had high
DOC. For P. rapida population
growth, DOC was not predictive
of chronic toxicity
Growth:
Effects on growth occurred at low
|jg/L concentration in L. stagnalis
in some of the tested waters. For
the snails, the greatest effects on
growth occurred with low-DOC
waters
C. dubia:
7-d-EC5os for
reproduction ranged from
20.1 to 573.4 |jg/L in
representative surface
waters of varying
chemistries. EC20S ranged
from 12.1 to 223.3 |jg/L.
P. rapida:
EC20 and EC50 ranged
from 3.2 to 103.3 and
10.6 to 154.9 |jg/L
dissolved Pb, respectively
L. stagnalis:
EC20S and ECsos for
growth ranged from 1.5 to
49.5 and 3.6 to
244.6 |jg/L dissolved Pb,
respectively, in the natural
waters
Esbauah et al.
(2012)
L. stagnalis'. 14-d
chronic toxicity test
for growth starting
with 7 to 10 dph
snails. Water
changes and food
replacement every
48 h
Rotifer For the Ca and Reproductive effects pH: Reproduction: For population size in Nvs et al
(Brachionus sei"ies: °f Pb (PbCh) 7he eq50 (based on filtered Pb) natural waters: (2016b)
calyciflorus) (nominal assessed in recently for population size differed by up
11-147
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Species
Concentration Exposure Method
Modifying
Factors
Effects on Endpoint
Effect Concentration
Reference
(published
since the
2013 Pb ISA)
concentration
range 46-
2,200 |jg Pb/L)
For DOC test
series: (nominal
concentration
range 100-
10,000 |jg Pb/L).
Total and filtered
Pb in each series
was quantified
hatched rotifers
exposed to Pb for 48-
hr (three
generations). Tests
were performed in
four series (varying
Ca, varying pH,
varying DOC, and
natural waters
collected from five
unpolluted
waterbodies in
different locations in
Europe)
ranged from 6.8
to 8.2 in natural
waters
DOC:
ranged from 3.2
to 31.5 in natural
waters
Temperature:
25°C
to 4.6-fold in the natural waters.
The highest toxicity was
observed in the synthetic
reference water. For the
modifying factor bioassays, both
population growth rate and
population size generally
decreased with increasing pH.
For DOC, toxicity expressed as
filtered Pb decreased
significantly with increasing DOC.
Ca was not protective
EC10 ranged from 52
(synthetic reference
water) to 231 |jg Pb/L
EC20 ranged from 75
(synthetic reference
water) to 336 |jg Pb/L
ECso ranged from 138
(synthetic reference
water) to 634 |jg Pb/L
(based on filtered Pb
concentration)
Snail
(Lymnaea stagnalis)
6, 12.5, 25,
100 |jg Pb/L
(analytically
verified)
Juvenile snail growth
was assessed in a
static renewal assay
over a 16-d period.
Primary focus of the
study was to
investigate possible
mechanisms of Pb
toxicity
Temperature: Growth:
23°C-25°C After 4 d, a moderate effect of
pH = 7.8 Pb on juvenile snail growth was
observed, severity of growth
inhibition increased after 8 d,
effects on growth occurred prior
to net Ca2+ flux in the snails,
inhibition of carbonic anhydrase
activity in the snail mantle also
showed no effect with Pb
EC20 (biomass) at 8 d of
exposure was 3.2 |jg L~1
Pb
Brix et al.
(2012)
EC20 (biomass) was
3.5 |jg L~1 Pb after 16 d of
exposure
11-148
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Snail
(Lymnaea stagnalis)
0.18 (control),
2.7 and
8.4 |jg Pb/L
(measured)
Newly hatched
juvenile snails were
exposed to Pb (as
Pb(N03)2 in Milli-Q
water) for 56-d in a
full lifecycle
assessment toxicity
test in a flow-through
system to assess
effects on survival,
growth and
reproduction (number
of egg masses, time
until first egg mass,
number of embryos
per egg mass). The
reproductive phase
started at day 32
(egg masses
appeared in the
control) and
continued till the end
of the study
Temperature:
24.8 ± 0.2°C
pH:
6.89 ± 0.06
DOC:
330 ± 7.02 |jM C
Survival:
Survival was significantly
decreased at the highest
concentration (8.4 |jg Pb/L) after
21-d exposure to the end of the
experiment
Growth:
Growth was significantly
decreased, even at the lowest
tested concentration (1 |jg Pb/L)
on day 28. By day 56, growth
remained significantly lower than
the controls in the 2.7 and
8.4 |jg Pb/L concentration;
however, snails exposed to
1.0 |jg Pb/L surpassed the
growth rates of the unexposed
snails. Inhibition of specific
growth rate at the 2.7 |jg Pb/L
exposure was observed during
the last week of the experiment.
Survival:
56-d chronic toxicity
NOEC = 2.7 |jg Pb/L
LOEC = 8.4 |jg Pb/L
Growth:
28-d
NOEC <1.0 |jg Pb/L
LOEC = 1.0 |jg Pb/L
Reproduction:
NOEC <1.0 |jg Pb/L
LOEC = 1.0 |jg Pb/L
Munlev et al.
(2013)
Reproduction:
For the number of egg masses
and time until first egg mass, the
NOEC <1.0 |jg Pb/L and
LOEC = 1.0 |jg Pb/L. No effects
on the number of embryos per
egg mass were observed at any
concentration tested. Individuals
exposed to the highest
concentration (8.4 |jg Pb/L) did
not reproduce during the lifecycle
test. Egg capsule and embryo
diameter after 7 d of
development were significantly
reduced at 2.7 |jg Pb/L (the
highest concentration in which
snails reproduced in the study)
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
Snail
(Lymnaea stagnalis)
Low |jg Pb/L
concentrations
(Pb was
measured in each
assay) EC values
are from
combined results
of Pb data from
multiple toxicity
tests
Series of 14-d
chronic toxicity
assays with single
metals (Pb as
Pb(NC>3)2) and binary
metal mixtures with
juvenile L. stagnalis
to assess effects on
relative growth rate.
Concentration-
response curves
were obtained by
compiling all the
single-metal toxicity
tests performed at
different times over a
2-yr period
Temperature:
25 ± 1°C
pH = 7.81 ± 0.20
DOC = 0.76 ±0.0
8 mg L"1
Alkalinity = 0.80 ±
0.05 mEq-L-1
Growth:
Inhibition of relative growth rate
was observed at low |jg Pb/L
concentrations, consistent with
other bioassays with L. stagnalis
14-d chronic toxicity:
EC10 = 4.0 |jg Pb/L
EC20 = 7.67 |jg Pb/L
ECso = 23.4 |jg Pb/L
Corresponding chronic
effect concentrations
based on free-ion activity:
ECio = 0.157 |jg Pb/L
EC20 = 0.320 |jg Pb/L
ECso = 1.08 |jg Pb/L
Cremazv et al.
(2018)
Mussel
(Hyridella australis)
(Hyridella depressa)
(Velesunio
ambiguus)
(Alathyria profuga)
(Cucumerunio
novaehollandiae)
(Hyridella drapeta)
Each acute
toxicity test
consisted of a
control and 10
concentrations,
which were based
on preliminary
range-finding
tests. Individual
test
concentrations
were not
reported.
Concentrations
were measured
Glochidia (larvae)
from gravid females
collected from two
different river
catchments in
southeastern
Australia were used
in the bioassays.
Four static tests were
conducted for each
mussel species and
exposure time (24,
48 or 72 hr) with
PbCI in reconstituted
freshwater. Viability
(as assessed by
valve closure) was
determined at the
end of the exposure
period
Temperature:
22 ± 1°C
pH 7.0 ± 0.2
Hardness
42 ± 4 mg CaCC>3
L"1
Alkalinity
22 ± 2 mg CaCC>3
L"1
Survival:
Pb sensitivity significantly
increased with each exposure
time and varied by species, with
greatest toxicity observed in C.
novaehollandiae
24-hr ECso (for valve
closure as a proxy for
viability) ranged from 176
to 274 |jg Pb
48-hr EC50 ranged from
102 to 165 |jg Pb/L
72-hr EC50 ranged from
65 to 110 |jg Pb/L
72-hr calculated NEC
ranged from 11 to
21 |jg Pb/L
Markich (2017)
11-150
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
Prawn
(Macrobrachium
nipponense)
0, 5, 10, 20, 40,
80, 160, 320 and
640 |jg Pb/L
(nominal values)
Acute toxicity
bioassay
12 |jg Pb/L,
25 |jg Pb/L
(measured)
Chronic growth
bioassay
For the 96-hr acute Temperature:
toxicity assay, 26 ± 1°C
juveniles were
exposed to Pb as Pb
acetate in semistatic pH 7.0-7.3
renewal (every 24 h)
conditions, survival
was assessed every
24 h. For the chronic
growth assay,
prawns were
exposed for 60 days
under the conditions
described for the
acute bioassay.
Prawns fed a
commercial diet twice
daily
dissolved oxygen
>6.5 mg/L
DOC: 190 pmol/L
Survival: LCso values decreased
overtime in the acute bioassay
from 24 to 96 h. Mortality was not
significantly affected by Pb
(12 |jg Pb/L or 25 pg Pb/L) in the
60-day chronic bioassay.
Growth: reductions in weight
gain and specific growth rate in
prawns exposed to 25 pg Pb/L,
but not in prawns exposed to
12 pg Pb/L
Acute toxicity test: Ding et al.
24-hr LCso = 646 pg Pb/L (2QM
48-hr
LCso = 250.6 pg Pb/L
72-hr
LCso= 175.6 pg Pb/L
96-hr
LCso= 131.3 pg Pb/L
60-d chronic bioassay:
Reduction in weight gain
observed at 25 pg Pb/L
(approx. 20% of the 96-hr
LCso)
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
Vertebrates
Zebrafish
(Danio rerio)
There was low
solubility of Pb in
the hard water;
the highest
concentration of
dissolved Pb
measured in hard
water was
590 |jg Pb/L at a
total Pb
concentration of
630 |jg Pb/L .
Highest
concentration
tested was
3,830 |jg Pb/L
(measured) in the
hard water while
the dissolved
fraction was
200 |jg Pb/L.
Newly hatched larvae
were tested in either
soft water or hard
water with Pb as Pb-
nitrate for 96-h.
Experiments were
conducted in six-well
culture plates with
10 mL water and 10
larvae per well.
Water was changed
every 24 h
Temperature:
28°C
Soft water
Hardness:
11.7 mg CaCOs/L
pH: 7.48
Na+ = 220 M,
K+ = 14 M
Ca2+ = 75 M
Mg2+ = 42 M
DOC = 0.9 mg/L.
Hard water:
hardness = 141 m
g CaCC>3/L
pH = 7.8
Na+ = 700 M
K+ = 38 M
Ca2+ = 1,350 M
Mg2+ = 336 M,
DOC = 3.5 mg/L
Survival: Pb was more toxic to
larvae in soft water than hard
water. No mortalities were
observed in the bioassays with
hard water even at the highest
tested concentration
Soft water:
96-hr LCso = 52.9 pg Pb/L
Hard water:
96-hr
LCso = >590 |jg Pb/L
Alsop and
Wood (2011)
11-152
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
Zebrafish
(Danio rerio)
2, 5, 10, 15, 20,
30 |jg Pb/L;
analytically
verified
concentration
Embryos/larvae were
exposed to Pb
acetate trihydrate
from 2 h
postfertilization (hpf)
embryos to 144 hpf
50% of the exposure
solution was
renewed daily
Temperature: Reproduction
28 ± 0.5°C No significant effect on
percentage of hatched larvae at
any of the tested concentrations
Growth
Significant increase in
prevalence of malformations at
30 |jg Pb/L compared with the
control
Zhu et al.
(2014)
Survival
Significant decrease in survival at
30 |jg Pb/L compared with the
control
Zebrafish
(Danio rerio)
5, 9.7,
19.2 |jg Pb/L;
measured
6-hpf embryos
exposed to Pb
acetate trihydrate
until 144-hpf. 50% of
exposure solution
was renewed daily
Temperature: Reproduction
28 ± 0.5 °C No significant difference on
hatching success rate at any of
the tested concentrations
Growth
No significant differences were
found for body length or body
weight at tested concentrations
compared with control
Zhu et al.
(2016)
Survival
No significant effect on survival
11-153
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
Zebrafish
19.3 |jg Pb/L
6-hpf embryos
Temperature:
Reproduction
Chen et al.
(Danio rerio)
exposed to Pb
28 ± 0.5°C
No significant difference on
(2016b)
acetate trihydrate
until 144-hpf. 50% of
exposure solution
was renewed daily.
Mortality rate,
malformation rate
(e.g., pericardial
edema and axial
spinal curvature) and
hatching success
recorded each day.
hatching success rate at
19.3 |jg Pb/L compared with
control.
Growth
No significant differences were
found for body length or body
weight at 19.3 |jg Pb/L compared
with control
After exposure, body
Survival
length and body
weight of each
No significant effect on survival
zebrafish larva was
at 19.3 |jg Pb/L
measured
Zebrafish
4.5, 9.6,
6-hpf embryos
Temperature:
Reproduction:
Zhao et al.
(Danio rerio)
18.6 |jg Pb/L
exposed to Pb
28.5°C
Hatching success rate
(2020)
analytically
verified
concentration
acetate tri hydrate
until 144-hpf. 50% of
exposure solution
was renewed daily.
significantly decreased in all
concentrations at 72 hpf
compared with control; this delay
in hatching rate also observed at
For each treatment,
96 hpf.
malformation,
survival rate and
hatching rate were
Survival
recorded at 24, 48,
Survival rate of Pb-exposed
72 and 96 hpf.
embryos at all tested
Additional behavioral
concentrations significantly lower
assays were
than controls at 96 hpf.
conducted at 144 hpf
11-154
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
rainbow trout
(Oncorhynchus
my kiss)
Waterborne-only
study:
4, 10, 20, 80, 240
and 800 |jg Pb/L
(nominal
concentration
reported,
concentrations
analytically
verified)
Waterborne, diet
and combined
exposure study:
0, 8.5, 20, 60 and
110 |jg Pb/L
(measured)
In waterborne
exposure to establish
LC/EC values,
juveniles (average
size = 2-4 g) were
exposed for 7 wk to
Pb as Pb-nitrate;
growth (weighed
weekly) and survival
were assessed at
various timepoints
including 96-h. In the
second study,
juvenile fish were
exposed for 7 wk via
waterborne Pb only,
dietary Pb only in the
form of live prey
(worms Lumbriculus
variegatus pre-
exposed for 28-d to
the same
concentration of Pb
as the fish) or
simultaneously to
waterborne and
dietary Pb
Temperature:
13°C
pH:
7.8-8.0
Hardness:
140 mg/L as
CaCC>3
DOC:
2.5 mg/L
Survival:
In the waterborne-only study to
establish LC/EC values, all fish in
the highest concentration tested
(800 |jg Pb/L) did not survive. In
the second study, survival in all
treatments (waterborne only,
dietborne only or combination)
and tested concentrations were
comparable to the control
(>90%).
Growth:
Waterborne Pb exposure had no
significant effects on specific
growth rate or biomass in either
experiment. In the dietary
combination experiment,
marginal (nonsignificant)
reductions were observed in the
dietborne and combined
exposures only at 110 |jg Pb/L
96-hr:
LCio = 304.3
H9
Pb/L
LC2o = 357.7
H9
Pb/L
LCso = 487.3
H9
Pb/L
Alsop et al.
(2016)
7-w:
LC10 = 55.6 |jg Pb/L
LC2o = 96.9 |jg Pb/L
LCso = 280.2 |jg Pb/L
11-155
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
rainbow trout
(Oncorhynchus
my kiss)
white sturgeon
(Acipenser
transmontanus)
Trout:
0, 10, 20, 40, 80,
160 |jg Pb/L
(nominal values)
Sturgeon:
0, 5.0, 10, 20, 40,
80 |jg Pb/L
(nominal values)
Measured
concentrations of
metals (not
provided) were
used for
calculation of
effect
concentration
A series of chronic
tests with two
lifestages (newly
hatched larvae and
approximately 1-mo-
old juveniles) of trout
and sturgeon were
conducted in
aqueous-only
exposure with Pb as
Pb-nitrate.
For trout: C1: 1-dph
larval trout in a 21-d
exposure; C2: 26-
dph juvenile trout in a
28-d exposure; CC:
1-dph larval trout in a
52-d exposure.
For sturgeon: C1: 2-
dph larval sturgeon in
a 25-d exposure C2:
27-dph juvenile
sturgeon in a 28-d
exposure; CC: 2-dph
larval sturgeon in a
53-d exposure. An
additional (C1-R) test
was conducted with
1-dph larval sturgeon
in a 24-d exposure
Trout:
Temperature:
12 ± 1°C
Hardness:
Approximately
100 mg/L as
CaC03,
Alkalinity:
approximately
90 mg/L as
CaCC>3
pH:
approximately
8.0
Sturgeon:
Temperature:
15 ± 1°C
Hardness:
Approximately
100 mg/L as
CaCC>3
Alkalinity:
approximately
90 mg/L as
CaCC>3
pH:
approximately
8.0
Growth/Survival
Note: Effect concentrations
reported in this study are based
on the most sensitive endpoint
(mortality, immobility, fish length
or biomass).
Trout: No acute effects observed
in larval or juvenile fish after 4-d.
Generally, trout were tolerant to
Pb concentration used in the
study
Sturgeon:
No mortality or immobilization of
newly hatched sturgeon was
observed by 4-d. The 53-d
exposures did not meet the test
acceptability criteria (due to
control mortalities); therefore,
there are no 53-d EC20S for the
survival. However, the EC20S
based on the length and weight
of surviving fish throughout the
53-d exposures were reported
Trout:
Acute 4-d EC50
C1 (larvae): >136 |jg Pb/L
C2 (juvenile):
>143 |jg Pb/L
CC (larvae) >136 |jg Pb/L
Chronic EC20
C1 (larvae 21-d)
>128 |jg Pb/L
C2 (juvenile 28-d)
>128 |jg Pb/L
CC (larvae 52-d)
>126 |jg Pb/L
Sturgeon:
Acute 4-d EC50
C1 (larvae): >55 |jg Pb/L
C2 (juvenile): >61 |jg Pb/L
CC (larvae) >55 |jg Pb/L
Chronic EC20
C1 (larvae 14-d)
>56 |jg Pb/L
C2 (juvenile 28-d)
>60 |jg Pb/L
CC (larvae 53-d)
>27 |jg Pb/L (note: low
control survival in this
experiment)
Wang et al.
(2014a)
11-156
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
white sturgeon
(Acipenser
transmontanus)
8 dph
Lab:
0.1, 0.8, 2.3,
19, 65, 210,
414 |jg Pb/L
6.4,
Columbia River:
0.2, 0.4, 1.4, 6.1,
17, 60, 191,
410 |jg Pb/L
40 dph
Lab:
0.1, 21, 46, 97,
208, 396, 809,
1610 |jg Pb/L
Columbia River:
0.3, 20, 37, 95,
192, 325, 799,
1685 |jg Pb/L
96-hr acute toxicity
assays conducted
with two lifestages (8
and 40 dph)under
static renewal
conditions with Pb as
Pb-nitrate in
laboratory water and
field-based tests with
Columbia River
water. The
laboratory- and field-
based tests were
conducted in parallel,
under the same
exposure conditions
and following the
same experimental
protocols. Water from
the Columbia River
was pumped into a
trailer retrofitted for
toxicity testing
Laboratory water
Temperature:
16 ± 0.9°C
PH
7.5 ± 0.2
Ca2+ to Mg2+
Ratio: -1.3:1
Columbia River
Water
Temperature:
16 ± 0.7°C
PH
7.7 ± 0.1
Ca2+ to Mg2+
Ratio: ~4:1
Survival
Fish exposed at 8 dph were more
sensitive than fish exposed at 40
dph. Fish exposed in lab water
were more sensitive than fish
exposed to Columbia River
water. There was a lack of
mortality observed in 8 dph fish
exposed to river water even at
the highest concentration tested.
8 dph
96-hr LCso= 177 pg Pb/L
(lab water)
96-hr
LCso = >410 |jg Pb/L
(Columbia River water)
40 dph
96-hr LCso = 528 pg Pb/L
(lab water)
96-hr
LCso= 1,556 pg Pb/L
(Columbia River water)
Vardv et al.
(2014)
11-157
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Species
Concentration Exposure Method
Modifying
Factors
Effects on Endpoint
Effect Concentration
Reference
(published
since the
2013 Pb ISA)
Asiatic toad
(Bufo gargarizans)
0, 9.85, 48.73,
97.69, 497.34 and
998.27 |jg Pb/L,
(measured)
corresponding to
0, 10, 50, 100,
500 and
1,000 |jg Pb/L,
nominal.
First larval stage
(Gosner stage 26)
tadpoles exposed to
Pb acetate in static
renewal (every 48 h)
solutions up to
Gosner stage 42
(forelimb emergence
starting at 31 to 35 d
depending on Pb
treatment group).
Tadpole growth and
developmental stage
assessed at day 10
and day 20.
Exposure continued
until day 60 to
determine mean
percent
metamorphosis
Temperature: Growth
~20°C On days 10 and 20, significant
increase reported in total tadpole
length and body mass at
50 |jg Pb/L. At Gosner
developmental stage 42
(metamorphic climax), snout-vent
length was significantly longer
than control in the 10 |jg Pb/L
treatment group. Snout-vent
length and total length were
significantly longer in tadpoles
exposed to 50 |jg Pb/L compared
with control. No statistically
significant difference in body
mass or tail length in any
treatment.
Survival
No mortality observed in control,
10, 50 or 100 |jg Pb/L during 60-
d exposure.
Yang et al.
(2019)
Dark-spotted frog
(Pelophylax
nigromaculata)
38.2, 79.3, 158.4,
318.7, 638.1,
1278.9 |jg Pb/L
analytically
verified
concentration
corresponding to
40, 80, 160, 320,
640,
1280 |jg Pb/L
nominal;
Embryos exposed to
Pb-nitrate in static
renewal assays from
heartbeat (Gosner
stage 19) to full
metamorphosis
(Gosner stage 46).
Chronic exposure
duration was up to
70 d
Temperature
19°C-25°C (room
temperature)
PH
7.04-7.69,
DO
6.8-7.3 mg/L
Hardness
249-258 mg
CaCOs/L
Growth
Growth was inhibited at higher
Pb concentrations; total
malformation rate increased
linearly with Pb concentration.
Survival
No significant effect on survival
at 40, 80, 160 or 320 pg Pb/L
Lowest threshold
concentration = 160 pg P
b/L for effects on
metamorphosis time,
body mass, snout-vent
length, and forelimb
length
Huang et al.
(2014)
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
Multiple
37 species and 32
genera of
invertebrates and
fish
(acute toxicity data
included in
derivation of
proposed updated
acute freshwater
quality criterion for
Pb)
15 species and 13
genera of
invertebrates and
fish (chronic toxicity
data included in
derivation of
proposed updated
chronic freshwater
quality criterion for
Pb)
Pb was
analytically
verified in all
studies
U.S. EPA guidelines
(U.S. EPA. 1985b)
were followed to
identify acceptable
studies. Water
chemistries over a
wide range of
conditions were
predicted from the
biotic ligand model.
Acute: All included
assays were
waterborne Pb
exposures reporting
48 to 96-hr ECsos.
The four lowest
genus mean acute
values (Hyalella,
Ceriodaphnia,
Gammarus and
Daphnia) and a total
of 32 genus mean
values were used to
determine a 50th
percentile critical
accumulation
concentration to
derive the proposed
acute criterion based
on U.S. EPA
methods
Acute toxicity end points
included survival,
immobilization, and loss of
equilibrium
The proposed updated acute
criterion is based on expanded
toxicity data sets and BLM
predictions that demonstrate the
influence of water hardness, used
in the calculation of the current
water quality criteria, is less
important as a modifying factor
relative to DOC.
Chronic toxicity endpoints
included survival, growth, and
reproduction
There is sufficient new chronic
toxicity data for Pb since the
1984 water quality criteria to
allow for direct determination of
criteria from toxicity data, rather
than the use of an acute-to-
chronic ratio.
Proposed Freshwater
Acute Water Quality
Criterion based on BLM of
North American surface
water chemistry
conditions ranged from
18.9 to 998 |jg Pb/L.
Proposed Freshwater
Chronic Water Quality
Criterion based on BLM of
North American surface
water chemistry
conditions ranged from
0.37 to 41 |jg Pb/L
Deforest et al.
(2017)
Chronic: Based on
EC20 values from
lifecycle tests in
freshwater
invertebrates as well
11-159
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Reference
Species Concentration Exposure Method Effects on Endpoint Effect Concentration (published
r r Factors r since the
2013 Pb ISA)
as partial lifecycle or
early lifestage tests
in fish. The four
lowest genus mean
chronic values
(Lymnaea, Philodina,
Hyalella,
Ceriodaphnia) and a
total of 13 genus
mean values were
used to identify a
chronic 5th percentile
waterborne Pb
concentration
following U.S. EPA
guidelines
Ca2+ = calcium ion; CaC03 = calcium carbonate; d = day; DOC = dissolved organic carbon; dph = days posthatch; DT = diatom + Tetramin; ECX = X% effect concentration;
hpf = hours postfertilization; K+ = potassium ion; LCX = X% lethal concentration; Mg2+ = magnesium ion; mo = month(s); Na+ = sodium ion; Pb = lead; Pb(N03)2 = lead nitrate;
wk = week(s); YCT = yeast, cereal leaves, and trout; yr = year(s).
11-160
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11.3.6
Freshwater-Community and Ecosystem Effects
Field studies in the 2006 Pb AQCD (U.S. EPA, 2006) and the 2013 Pb ISA (U.S. EPA, 2013)
report reductions of species abundance, richness, or diversity, particularly in benthic macroinvertebrate
communities coexisting with multiple metals where the sources of Pb were from mining or urban
effluents. Changes to aquatic plant community composition have been observed in the presence of
elevated surface water Pb concentrations. Additionally, field studies have linked Pb contamination to
reduced primary productivity and respiration, and to altered energy flow and nutrient cycling. In the 2013
Pb ISA (U.S. EPA, 2013), the body of evidence was sufficient to conclude there is a likely to be causal
relationship between Pb exposure and freshwater-community and ecosystem effects. Studies reviewed in
that document noted ecological effects on invertebrate communities can occur at environmental Pb
concentrations lower than those required to affect plant communities. High sediment Pb concentrations
were linked to shifts in amphipod communities inhabiting plant structures, and potentially to alterations in
ecosystem nutrient processing. Although the presence of Pb is associated with shifts in biological
communities, this metal rarely occurs as a sole contaminant in natural systems, making the contribution of
Pb to the observed effects difficult to ascertain. New information on the effects of Pb at the population,
community, and ecosystem levels is reviewed below.
Several studies reviewed here reported negative associations between sediment Pb concentration
and invertebrate community composition. A series of studies conducted in Caddo Lake, Texas has further
elucidated the effects of Pb on benthic macroinvertebrate communities and Pb as a modifying factor in
leaf-litter decomposition. Caddo Lake is a shallow, eutrophic lake which neighbors a superfund site
(Longhorn Army Ammunition Plant, Texas). Oguma and Klerks (2015) found evidence that Pb
contamination may affect leaf-litter decomposition in the lake. Litter decomposition (relative change in
dry weight of American lotus [Nehimbo lutea] leaves deployed in litter bags) was determined after
30 days at sites spanning a gradient of sediment Pb concentration. Sediment Pb concentration in Caddo
Lake ranged from 4.3 to 148.9 mg Pb/kg, with some sites exceeding the Probable Effects Concentration
for sediment (128 mg Pb/kg). In a principal component analysis, total sediment Pb and sediment
porewater Pb were positively correlated, and benthic macroinvertebrate abundance was negatively
correlated with sediment Pb concentration and porewater Pb concentration. The authors suggested that the
combination of sediment Pb content and decreased macroinvertebrate abundance, among other untested
factors, may lead to reduced leaf-litter decomposition in Caddo Lake.
In a study on sediment macroinvertebrates in Caddo Lake, sediment Pb concentration was
negatively correlated with the diversity and abundance of benthic macroinvertebrates although amphipod
sensitivity to Pb and Cu was unrelated to sediment Pb and Cu concentrations (Oguma and Klerks, 2020).
Using a univariate approach between benthic community metrics and heavy-metal concentrations, the
benthic macroinvertebrate abundance, family richness, and Shannon H' Index were negatively correlated
with sediment Pb concentrations. Although this study provides correlational evidence that Pb sediment
11-161
-------
concentration affects benthic macroinvertebrate community structure, % sand/clay content, % OM, and
Cu sediment concentration among other principal components are correlated with benthic
macroinvertebrate community metrics. A sensitive amphipod (H. azteca) was exposed to sediment, and
reproduction, survival and growth were assessed at 28, 35, and 42 days. The survival (28, 35, and
42 days), reproduction (35 and 42 days) and growth (42 days) of H. azteca were not affected by Pb
sediment concentration.
Crayfish density was negatively correlated with sediment Pb concentration in the Old Lead Belt
mining district in Missouri where Pb-Zn mining occurred from the 1700s to the 1970s (Allert et al..
2013). Parts of the district were designated as U.S. EPA Superfund sites. To test whether benthic
macroinvertebrate, fish, and crayfish communities differed along Pb and other heavy-metal gradients in
the Big River, benthic fish, crayfish, macroinvertebrates, sediment, and surface waters were sampled from
riffles from eight sites (two reference sites where no mining activities occurred, two mining sites with
high contamination, and four sites downstream of the mining sites with slightly lower contamination).
The density of fish including sculpins (Cottiis spp.), darters (Etheostoma spp. And Percina spp.), and
madtoms (Noturus), and crayfish (Orconectes spp.) was estimated in situ. Individuals of the Missouri
saddled darter (Etheostoma tetrazonum) and golden crayfish {(). luteus) were collected and used for metal
analyses. Additionally, an in situ toxicity test on juvenile O. luteus and O. hylas was conducted at the two
reference sites and two mining sites over 56 days, and the growth and survival of crayfish were assessed
at the end of the test. Surface water Pb concentrations were lowest at the reference sites
(0.06 ± 0.01 |ig Pb/L, mean ± S.D.) and highest at the mining sites (7.85 ± 1.63 |ig Pb/L). Sediment Pb
concentrations followed the same pattern, with the lowest concentrations at the reference site
(12.5 ±2.1 mg Pb/kg dry weight), followed by the downstream sites (710 ± 530 mg Pb/kg dry weight)
and the highest concentrations at the mining sites (1170 ± 467 mg Pb/kg dry weight). Pb in the sediment
at the mining and downstream sites was significantly higher than the Probable Effects Concentration for
sediment derived by (MacDonald et al.. 2000) (128 mg Pb/kg dry weight). Pb concentration in detritus
was significantly lower in reference sites compared with mining sites. Moving up the food web, Pb
concentration in macroinvertebrates was lower in reference sites than in mining sites
(12.7 ± 4.4 mg Pb/kg dry weight for reference sites and 720 ± 276 mg Pb/kg dry weight for mining sites,
respectively). Similarly, in two different larval species of caged crayfish ((). luteus and O. hylas), Pb
concentration was lower in reference sites compared with the mining site. Field-collected adult O. luteus
Pb concentration followed the same pattern, reference Pb < downstream Pb < mining sites Pb. Pb
concentration in E. tetrazonum was highest in the mining sites (mean ± S.D., 66.8 ± 7.3 mg Pb/kg dry
weight), followed by the downstream sites (44.7 ± 14.4 mg Pb/kg dry weight) and the reference sites
(0.55 ± 0.14 mg Pb/kg dry weight). Orconectes luteus carapace length (mm) was significantly negatively
correlated with sediment Pb concentration, surface water Pb concentration, and Orconectes luteus Pb
concentration. Sediment Pb concentration was significantly negatively correlated with crayfish density
(number of crayfish x m 2). Surface water Pb concentration was significantly negatively correlated with
fish density and crayfish density. Although whole-body E. tetrazonum Pb concentration was not
significantly correlated with fish density or crayfish density, O. luteus whole-body Pb concentration was
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significantly negatively correlated with crayfish density. Benthic fish density (number of benthic
fish x m 2) and crayfish density (number of crayfish/ m 2) were significantly reduced under high Pb
Probable Effects Quotient values, defined as the Probable Effects Concentration divided by the total
recoverable metals in the sediment.
In a field study, bioaccumulation of Pb and Cd in the common reed (Phrctgmites australis) was
correlated with the density of periphyton in aquatic ecosystems in Greece (Obolewski et al.. 2011). Forty-
five reed sampling sites around Greece included saltwater lagoons, bays, freshwater lakes, dam reservoirs,
irrigation and wastewater canals, and a river representing a gradient of hydrological parameters, salinity,
water movement and contaminants. The concentrations of Pb in P. australis shoots varied among
ecosystems and seasons, but most concentrations were between 19 and 21 mg Pb/kg for all sites and
seasons. Using a redundancy analysis, biplot scores indicated that Pb was negatively correlated with
Oligochaeta. Cyanophyta was found in sites with higher concentrations of Pb, Cd and Cu (and correlated
metals Zn, Ni, Co, and Fe). Scendesmns were found in sites with lower concentrations of Pb and Mn (and
correlated with Zn, Ni, Co, Fe).
The Pb gradient was not strongly correlated with shifts in aquatic insect diversity in Swedish
lakes and ponds near an abandoned Zn-Pb mine (Lidman et al.. 2020). The most important variables
associated with larval insect community composition were bioavailable Zn, sediment Zn, bioavailable Pb,
Ca, NO3, and NH4. For adult macroinvertebrate communities, bioavailable Pb, and sediment Pb were not
statistically significant. In the analyses of larval and adult aquatic insect communities, sediment Pb was
negatively correlated with community structure, while bioavailable Pb was positively correlated with
community structure.
In summary, new observational and experimental studies published since the 2013 Pb ISA (U.S.
EPA. 2013) reported either negative, positive, or null associations between sediment or porewater Pb
concentration and community and ecosystem effects. Specifically, benthic macroinvertebrate abundance
and leaf-litter decomposition were negatively correlated to sediment Pb concentrations in freshwater lakes
(Oguma and Klerks. 2015). Macroinvertebrate community composition was found to be sensitive to mild
Pb contamination in a freshwater lake (Oguma and Klerks. 2020). Crayfish and fish density was
negatively correlated to surface water Pb concentrations and sediment concentrations for crayfish in a
river system (Allert et al.. 2013). Pb accumulated in reeds were found to be negatively, positively, or not
correlated with abundance of some periphyton families (Obolewski et al.. 2011) Finally, larval and adult
insect community structures were affected by natural gradients of Pb in a lake system (Lidman et al..
2020).
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11.4
Saltwater Ecosystems
11.4.1 Summary of New Information on Effects of Pb in Saltwater Ecosystems
and Causality Determination Update Since the 2013 Pb ISA
Historically, the effects of Pb were less well characterized in saltwater biota compared with
freshwater biota. In field studies of coastal and marine saltwater ecosystems it is difficult to attribute
observed effects solely to Pb due to the presence of other stressors and highly variable conditions which
influence Pb speciation and toxicity in these environments. Furthermore, the portion of Pb from
atmospheric sources is usually not known. Most of the information on Pb effects on saltwater organisms
are from laboratory-based studies. Fewer toxicity bioassays have been conducted on saltwater plant and
algal species compared to freshwater species, and the observed effects generally occurred at
concentrations that greatly exceeded reported concentrations of Pb from coastal waters (Table 11-1).
Evidence in the 2013 Pb ISA was inadequate to infer causality relationships between Pb exposure and
effects on physiological stress, growth, survival, and reproduction in saltwater plants and algae (U.S.
EPA, 2013). In the 1977 Pb AQCD and the 1986 Pb AQCD, there were no studies that reported the
effects of Pb in saltwater invertebrates. In the 2006 AQCD, few effects were noted in saltwater
invertebrates including gender differences in sensitivity to Pb in copepods, increasing toxicity of Pb with
decreasing salinity in mysids and effects on embryogenesis in bivalves (U.S. EPA, 2006). In the 2013 Pb
ISA, available evidence was sufficient to be suggestive of a causal relationship between Pb exposure and
the endpoints of physiological stress, hematological effects, and reproduction for saltwater invertebrates
(U.S. EPA, 2013). Evidence for effects on neurobehavior, growth and survival in saltwater invertebrates
and vertebrates, as well as effects on ecological populations and communities, was concluded to be
inadequate to infer a causality relationship.
For many of the endpoints for saltwater biota (Table 11-7), evidence remains inadequate to assess
causality. For other endpoints, new evidence continues to support, or expands somewhat, the basis for the
causality determination in the 2013 Pb ISA. For suborganism-level endpoints, evidence was suggestive of
a causal relationship between Pb exposure and physiological stress in saltwater invertebrates in the 2013
Pb ISA, and this remains the case. There is very little new evidence for hematological effects of Pb in
saltwater invertebrates, which, at the time of the 2013 Pb ISA, was suggestive of, but not sufficient to
infer, a causal relationship (U.S. EPA, 2013). Evidence for hematological effects in previous AQCDs and
the 2013 Pb ISA were primarily from field monitoring studies of marine bivalves using ALAD as a
biomarker for Pb exposure and correlated ALAD inhibition to increased Pb tissue content. For the
organism-level endpoints of neurobehavior and growth effects associated with Pb exposure, there is
inadequate experimental evidence to assess causality for saltwater species.
Since the 2013 Pb ISA, there is additional research for saltwater organisms that supports a change
in causality determinations for some endpoints. Several newer studies quantify Pb in exposure media and
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report effects on endpoints at lower concentration than previously observed for saltwater biota. The
increased availability of studies that report analytically verified concentrations have enabled updated
estimates of effects criteria. For example, an increase in toxicological data for saltwater organisms over
the last several years and the availability of studies that analytically verified Pb exposure concentration
has led to a study proposing updates to the acute and chronic AWQC for Pb (Church et al.. 2017). For the
acute criterion, the proposed update of 100 |ig Pb/L is less than the current acute criterion of 210 (ig Pb/L
due to more recent toxicity data from relatively sensitive early lifestages of Echinodermata and Mollusca.
In the 2013 Pb ISA, the evidence at that time for Pb effects on the survival of saltwater
vertebrates was inadequate to infer a causal relationship with Pb exposure (U.S. EPA, 2013). New
evidence (Section 11.4.5) is limited to laboratory-based bioassays in a few fish species. Toxicity data for
other saltwater vertebrates remains lacking. Several recent chronic bioassays conducted with early
lifestages of three saltwater fish species reported NOEC in the range of 11-14 |ig Pb/L (Table 11-7).
Furthermore, Pb in these bioassays was analytically verified. In the larval fish topsmelt (Atherinops
afflnis), LC50 = 15.1 (ig Pb/L and NOEC <13.8 |ig Pb/L were obtained at a salinity of 14 ppt (Reynolds et
al., 2018). Calculated chronic values for additional saltwater fish species that are consistent with the range
reported above include grey mullet (Miigil cephahis) fingerling survival and Tiger perch (Terctpon
jarbuct) fingerling survival (Hariharan et al., 2016). Based on these new chronic studies in saltwater fish,
the evidence is suggestive of, but not sufficient to infer, a causal relationship between Pb exposure
and saltwater vertebrate survival.
In the 2013 Pb ISA the evidence was concluded to be suggestive of, but not sufficient to infer, a
causal relationship between Pb exposure and reproduction and developmental effects in saltwater
invertebrates (U.S. EPA, 2013). Endpoints reported in the previously available studies included a delay in
the onset to reproduction (amphipod Elasmopus Ictevis) (Ringenary et al„ 2007), impaired larval
development (Wang et al., 2009) and embryogenesis inhibition (Wang et al., 2009; Beiras and Albentosa,
2004) in bivalves and a decrease in the fertilization rate of eggs (marine polycheate annelid Hydroides
elegctns) (Gopalakrishnan et al., 2008). Since the 2013 Pb ISA, the evidence base for Pb effects on
reproductive and developmental endpoints in saltwater invertebrates has expanded, primarily due to
multiple new embryo-larval developmental assays in Mollusca and Echinodermata (Section 11.4.5 and
Table 11-7). Several of these acute exposure bioassays analytically verified the concentration of Pb at
which effects were observed (Markich, 2021; Romero-Murillo et al„ 2018; Nadella et al„ 2013) and
reported effects at lower concentrations than those reported in the 2013 Pb ISA. The 48-hour EC10 larval
development in the mussels Mytilus trossulus and Mytihis galloprovincialis, was 9 and 10 |ig Pb/L
respectively, and 72-hour EC10 was 19 |ig Pb/L in the sea urchin Strongylocentrotus piirpiiratus (Nadella
et al., 2013). In the scallop Argopectenpurpuratus, there was a 48-hour EC50 = 44 |ig Pb/L for abnormal
larval development (Romero-Murillo et al„ 2018). These effects concentrations are comparable to those
reported for larval developmental assays from two species of oysters Magallanct gigas (48-hour
EC50 = 49.5 |ig Pb/L, 48-hour NEC = 9.9 |ig Pb/L) and Saccostrea glomerate/ (48-hour
EC50 = 52.1 |ig Pb/L, 48-hour NEC = 10.1 |ig Pb/L) (Markich, 2021). Considering the coherence of
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reproductive and developmental effects of Pb across species, observations in saltwater invertebrates are
consistent with terrestrial and freshwater invertebrates (both ""causal" in the 2013 Pb ISA) As a result of
the newly available evidence since the 2013 Pb ISA, the evidence is sufficient to conclude there is
likely to be a causal relationship between Pb exposure and reproductive and developmental effects
in saltwater invertebrates.
For community and ecosystem effects, evidence was inadequate in the 2013 Pb ISA to assess
causality between Pb exposures and the alteration of species richness, species composition and
biodiversity in saltwater ecosystems. Reduced species abundance and the biodiversity of protozoan and
meiofauna communities were observed in laboratory microcosm studies with marine water and marine
sediments reviewed in the 2006 Pb AQCD, as summarized in Table AX7 2.5.2 (U.S. EPA, 2006). In the
2013 Pb ISA, there were a few additional studies including effects on community structure and nematode
diversity that were altered in a microcosm study with marine sediments (Mahmoudi et al., 2007). Since
then, new experimental and observational studies have examined the relationship between Pb in sediment
and microbial abundance and/or diversity (Section 11.4.4.1), as well as Pb associations with saltwater
foraminiferal communities (Section 11.4.6). Several of the benthic foraminifera studies reported effects
on community richness, diversity, and abundance. In other studies with foraminifera, there were changes
in the abundance of certain taxa associated with Pb, but not diversity metrics. Considering the new
evidence, Pb quantified in sediment is a factor that explains variations in microbial diversity and
foraminiferal species distributions and abundance in a variety of distinct geographic locations. In these
studies, Pb was often correlated with other heavy metals.
These effects observed in saltwater biota are coherent with the observed community and
ecosystem-level effects of Pb in terrestrial and freshwater environments, which were reported as "likely
causal" in the 2013 Pb ISA (U.S. EPA, 2013). In addition to the available studies assessing Pb effects on
saltwater communities, primarily foraminifera, the effects of Pb on reproduction in sensitive saltwater
invertebrates and possible effects on survival in early lifestages of some saltwater vertebrates, especially
when considered cumulatively, could affect populations as well as community and ecosystem structure
and function. Population, community, or ecosystem-level studies are typically conducted at sites that have
been affected by multiple stressors (several chemicals alone or combined with physical or biological
stressors), which increase the uncertainty of attributing the observed effects to Pb. Therefore, for saltwater
the evidence is suggestive of, but not sufficient to infer, a causal relationship between Pb exposure
and community and ecosystem effects.
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Table 11-6 Updated causality determinations for Pb in saltwater organisms
and ecosystems
Level
Effect
Saltwater3
2013 Pb ISAb
2024 Pb ISAC
Community and
Ecosystem
Community and Ecosystem Effects
Inadequate
Suggestive
Reproductive and Developmental Effects
- Plants
Inadequate
Inadequate
Reproductive and Developmental Effects
- Invertebrates
Suggestive
Likely Causal
Population-
level
Reproductive and Developmental Effects
- Vertebrates
Inadequate
Inadequate
Growth - Plants
Inadequate
Inadequate
Endpoints
Organism-level
Responses
Growth - Invertebrates
Inadequate
Inadequate
Growth - Vertebrates
Inadequate
Inadequate
Survival - Plants
Inadequate
Inadequate
Survival - Invertebrates
Inadequate
Inadequate
Survival - Vertebrates
Inadequate
Suggestive
Neurobehavioral Effects - Invertebrates
Inadequate
Inadequate
Neurobehavioral Effects - Vertebrates
Inadequate
Inadequate
Hematological Effects - Invertebrates
Suggestive
Suggestive
Hematological Effects - Vertebrates
Inadequate
Inadequate
Suborganismal
Responses
Physiological Stress - Plants
Inadequate
Inadequate
Physiological Stress - Invertebrates
Suggestive
Suggestive
Physiological Stress - Vertebrates
Inadequate
Inadequate
Conclusions were based on the weight of evidence framework for causal determination in Table II of the ISA Preamble (U.S. EPA.
bEcological effects observed at or near Pb concentrations measured in sediment and water in Table 6-2 of the 2013 Pb ISA were
emphasized, and studies generally within one to two orders of magnitude above the reported range of these values were
considered in the body of evidence for saltwater (Section 6.4.21) (U.S. EPA. 2013).
°Changes from the 2013 Pb ISA are indicated as bolded text.
The 2013 Pb ISA concluded that the body of evidence was suggestive of a causal relationship
between Pb exposure and physiological stress, hematological effects, and reproductive and developmental
effects in saltwater invertebrates (Table 11-6). Evidence was inadequate at the time to assess causality for
additional effects in saltwater invertebrates and for marine algae and vertebrates. Key uncertainties from
the last review for saltwater ecosystems included the uncertainties associated with generalization of
effects observed in controlled laboratory studies to conditions in coastal environments where many
modifying factors affect Pb bioavailability and toxicity. In general, Pb toxicity to marine or estuarine
plants, invertebrates and vertebrates was less well characterized than toxicity to Pb in freshwater systems
in the 2013 Pb ISA due to an insufficient quantity of studies on saltwater organisms. Specifically, there
was a lack of chronic toxicity data, and relatively few studies reported analytically verified Pb
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concentration in the experimental media. Information regarding the contribution of atmospheric Pb to
total Pb in coastal environments was sparse. This was attributed to multiple sources of Pb, confounding
effects of transport from terrestrial and freshwater systems and the lack of studies connecting the air
concentration of Pb and saltwater ecosystem exposure.
Studies published since the 2013 Pb ISA (literature cutoff for inclusion in the 2013 Pb ISA was
September 2011) that characterized bioavailability, uptake, bioaccumulation, and effects of Pb in
saltwater ecosystems or that decreased uncertainties identified in the prior NAAQS review of this criteria
air pollutant are presented throughout the following sections. Saltwater ecosystems considered encompass
a range of salinities from just above that of freshwater (<1 ppt) to that of seawater (generally described as
35 ppt). Coastal ecosystems may receive Pb from multiple sources such as contributions from
atmospheric deposition and via inputs from terrestrial systems including runoff and riverine transport
(Appendix 1: https://assessments.epa.gov/isa/document/&deid=359536). Habitats associated with coastal
areas include salt marshes, estuaries, shallow open waters, sandy beaches, mud and sand flats, rocky
shores, oyster beds, coral reefs, mangrove forests, river deltas, tidal pools, and seagrass beds (U.S. EPA,
2016). Estuaries, where freshwater inflows gradually mix with salt water, are dynamic, heterogeneous
environments characterized by gradients of salinity. Salinity is one of the modifying factors affecting Pb
speciation in coastal systems, and changes in salinity affect the ionic strength of the water
(Section 11.4.2). The Pb2+ ion, which is the most bioavailable form of Pb, is not common in seawater;
rather, Pb primarily exists as a carbonate complex and to a lesser extent as a chloride complex (Church et
al.. 2017; Millero et aL 2009).
Brief summaries of conclusions from the 1977 Pb AQCD (U.S. EPA, 1977), the 1986 Pb AQCD
(U.S. EPA. 1986). the 2006 Pb AQCD (U.S. EPA. 2006) and the 2013 Pb ISA (U.S. EPA. 2013) are
included where appropriate. Recent research on the bioavailability and uptake of Pb into saltwater
organisms including plants, invertebrates and vertebrates is presented in Section 11.4.2. Section 11.4.3
covers environmental concentrations of Pb in saltwater biota and ecosystems in the United States at
different locations and overtime. The toxicity of Pb to marine flora and fauna including growth,
reproductive and developmental effects (Section 11.4.4) is followed with data on exposure and the
response of saltwater organisms (Section 11.4.5). Responses at the community and ecosystem levels of
biological organization are reviewed in Section 11.4.6.
11.4.2 Factors Affecting Bioavailability, Uptake and Bioaccumulation, and
Toxicity in Saltwater Biota
The environmental fate processes affecting Pb in the marine environment are distinct from
freshwater environments. Pb speciation in seawater is a function of chloride concentration and the
primary species are PbCl, PbCO . PbCl. and PbCl while in freshwater, Pb2+ is the predominant species
(U.S. EPA. 2013). PbCl is poorly soluble and will tend to precipitate and be found in bottom sediments.
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The generally high pH and salinity of marine systems, and in some cases high DOC and particulate matter
as well in estuarine waters, create conditions in which the percentage of free Pb2+ tends to be very low
(Appendix 1, Section 1.3.3.1.2).
Factors affecting bioavailability of Pb to saltwater organisms are many of the same factors
affecting bioavailability to freshwater biota (Section 11.3.2), notably OM, particulate matter, other
minerals, DO, and pH. Other factors, such as salinity, play a greater role in Pb fate, transport, and
bioavailability in saltwater systems, especially in dynamic estuarine environments characterized by
gradients of salinity. Marine environments are characterized by higher levels of ions, such as Na+, Ca2+,
and Mg2+, which compete for potential binding sites on biotic ligands such as gills, thereby generally
reducing the effective toxicity of metal ions as compared to freshwater environments (U.S. EPA, 2013)
(See also Appendix 1, Section 1.3.3.1.2). Since the 2013 Pb ISA, there is additional information
(summarized below) on these chemical factors which can be quantified and directly related to toxicity.
Studies have further explored the effects of varying DOM composition and changing pH on Pb uptake
and bioaccumulation in saltwater biota. Other factors that affect the uptake and toxicity of Pb, such as
biological adaptations by organisms, are more difficult to link quantitatively to toxicity. As discussed in
previous U.S. EPA reviews of Pb, species differences in metabolism, sequestration, and elimination rates
have been shown to control the relative sensitivity and vulnerability of exposed organisms and influence
the potential for effects on survival, reproduction, growth, metabolism, and development. Diet and
lifestage at the time of exposure also contribute significantly to sensitivity and vulnerability in
populations and communities. The 2006 Pb AQCD (U.S. EPA, 2006) reviewed the effects of genetics,
age, and body size on Pb toxicity. While genetics appears to be a significant determinant of Pb sensitivity,
the effects of age and body size are complicated by environmental factors that alter the metabolic rates of
saltwater organisms. Literature reviewed in the 2013 Pb ISA corroborated these findings and discussed
seasonal physiological changes and lifestage as important determinants of differential sensitivity to Pb.
11.4.2.1 Dissolved Organic Matter
In seawater, DOM is a major factor controlling bioavailability of Pb (U.S. EPA, 2013). Studies
reviewed in the 2013 Pb ISA showed that different components of DOM have different effects on Pb
bioavailability in marine systems. Increasing humic acid concentrations increased Pb uptake by mussel
gills and increased toxicity to sea urchin (Paracentrotus lividus) larvae (Sanchez-Marin et al„ 2007),
while in contrast, fulvic acid reduced Pb bioavailability (Sanchez-Marin et al., 2011). Continuing their
research in a study published after the 2013 Pb ISA Sanchez-Marin and Beiras (2012) observed that more
soluble DOM (fulvic acids and DOM extracted from the Suwannee River) also increased the
bioavailability and toxicity of Pb to sea urchin embryos, although not to the same extent as humic acid.
Furthermore, the experimental evidence suggests that the mechanisms by which DOM enhances Pb
uptake and toxicity implies direct contact of the organic compounds with the plasma membrane. In
another study examining the effects of different forms of DOM, Tang et al. (2020) observed that the
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bioaccumulation of Pb in saltwater shrimp was likely affected by the quality of OM; with more
autochthonous OM present, there was less bioaccumulation compared with the levels in winter months
when more allochthonous OM is present. Additionally, because the ingestion of DOM bound to metals is
the major route of entry for metals, this suggests that the allochthonous OM may have a greater
percentage of functional groups that bind Pb (e.g., fluorophores).
Several studies published since the 2013 Pb ISA have explored the protective effects of different
types of OM by quantifying enzymatic activity and oxidative response in saltwater invertebrates.
Nogucira et al. (2018) examined the toxicity of Pb alone and in combination with natural OM (NOM)
from different sources (allochthonous, autochthonous, and mixed) on larvae of the Canadian native bay
mussel (Mytilus trossulus). With 48-hour exposure to Pb alone (20 |ig Pb/L) there was an increase in
carbonic anhydrase activity and lipid peroxidation. Various NOMs did not protect against Pb toxicity, and
lipid peroxidation increased significantly with some types of NOM. A parallel study conducted on the
invasive Mediterranean mussel (Mytilus galloprovincialis) (Nogueira et al., 2017) also showed that
various sources of NOM differentially induced increases of enzyme activities and oxidative stress to a
greater extent than Pb alone; however, M. galloprovincialis was less sensitive than native M. trossulus
overall. In these studies, no protective effects of NOM were observed. The interaction of NOM with
metals is influenced by the source and composition of NOM, and some forms of NOM may exert a
sublethal response independently. In a series of bioassays, Nadella et al. (2013) assessed the influence of
DOM on the embryo development of two mussels, M. galloprovincialis and M. trossolus, and the pacific
purple sea urchin (S. purpuratus). Addition of DOM from a freshwater source and a seawater source
decreased the toxicity of Pb to embryos of the mussels compared with toxicity tests in 100% seawater.
However, there was no concentration-dependent relationship with increasing addition of DOM.
Unexpectedly, DOM exacerbated Pb toxicity in 48-hour embryo toxicity tests with S. purpuratus. In the
absence of Pb, one of the DOM sources resulted in 100% mortality of S. purpuratus embryos. The
authors speculated that this is a species-dependent response, attributable to DOM interaction with the
epithelial interface.
11.4.2.2 pH
The importance of pH in the speciation of Pb in saltwater environments and as a modifying
factor of Pb toxicity was previously reported (U.S. EPA, 2013, 2006). Several additional studies
published since the 2013 Pb ISA further describe pH effects on Pb uptake and toxicity in saltwater
organisms. A decrease in pH under the scenario of increasing ocean acidification may lead to additional
bioavailable Pb (Pb2+) in marine environments (Figure 11-5) and associated toxic effects on biota as
reviewed in Ivanina and Sokolova (2015). Belivermis et al. (2020) demonstrated that a decrease in pH
(from 7.94 to 7.16) resulted in a significant increase in 21"Pb in the soft tissues, but not the shells, of blue
mussels (M ednlis) after a 9-day exposure. Pb uptake in mussels was highly variable, likely due to the
variability of the physiological status of individual mussels. The lower Ca2+ in acidified seawater can
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make Pb2+ more available to mussels due to decreased competition, and the lower pH means a higher
partial pressure of CO2, which can result in decreased biomineralization that may facilitate the uptake
of Pb.
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Figure 11-5 Main forms of Pb in seawater as a function of pH at 25°C and
salinity of 35 ppt.
11.4.2.3 Salinity
In marine and estuarine systems, salinity is an important factor influencing the speciation of
metals and subsequent bioavailability (de Sousa Machado et al„ 2016; Wright, 1995). Generally, an
increase in salinity reduces the bioavailability of metals by increasing complexing with chloride and
carbonate ions and decreasing the amount of Pb2+ (U.S. EPA, 2006; Wright, 1995). New information
published since the 2013 Pb ISA further characterizes the bioavailability of Pb under different salinity
levels. For coastal sediments, Liu et al. (2019a) observed that the bioavailability of Pb at 35 ppt salinity
was sequentially higher than that at salinity levels of 25 ppt and 15 ppt. When salinity was 35 ppt, the
bioavailable fractions of Pb in surface sediments increased by 20.38% compared with Pb at a salinity of
15 ppt. However, it was found that excess dissolved phosphate resulted in the precipitation of Pb3(P04)2,
which was spurred on by the increased bioavailability of Pb. In tropical estuary wetlands, Chu et al.
(2015) found that increased salinity can increase Pb mobility. This is due to Pb being transformed
primarily into exchangeable and reducible fractions at higher salinity, making Pb more bioavailable. The
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exchangeable Pb fraction increased and the oxidizable fraction of Pb and carbonate bound fraction
decreased with increasing salinity.
In a study reviewed in the 2006 AQCD, Verslvcke et al. (2003) exposed the estuarine mysid
Neomysis integer to individual metals, including Pb, and metal mixtures under changing salinity. At a
salinity of 5%, the reported LC50 for Pb was 1140 |ig/L (95% CL = 840, 1440 (.ig/L). At an increased
salinity of 25%, the toxicity of Pb was substantially reduced (LC50 = 4,274 |ig/L [95% CL = 3,540,
5710 |ig/L|). The reduction in toxicity was attributed to increased complexation of Pb2+ with CI" ions.
Studies published since the 2013 Pb ISA have further considered salinity as a modifier of Pb uptake and
toxicity in saltwater invertebrates. The relationships between tissue concentration of Pb and inorganic
cations (Na+, Mg2+, K+, and Ca2+) were assessed in the Hong Kong oyster (Crassostrea hongkongensis) at
four different salinities at a single Pb concentration (3 |ig Pb/L, nominal) under laboratory conditions (Yin
and Wang. 2017). All four cations were negatively correlated with trace metal uptake by oysters; the
tissue concentration of Pb was lower at higher salinities during the 6-week exposure (due to decreasing
free-ion concentration of Pb at higher salinity). For the rotifer 1'males similis, exposed nominally to Pb
(13, 25, 50, 100 (ig Pb/L) in 5-day chronic reproductive toxicity tests conducted at four salinity conditions
(5, 15, 25 and 35 ppt), population density was highest at the lowest salinity, and toxicity increased with
increasing Pb concentration (Rebolledo et al.. 2021). As salinity increased, population density decreased
in all treatments and the control; however, across all salinities, the population growth rate was lowest at
100 |ig Pb/L (the highest tested concentration). In contrast, embryo development assays in larval mussels
(bay and Mediterranean) and pacific purple sea urchins conducted at two salinities (33 ppt and 21 ppt)
reported no effect of salinity on Pb toxicity (Nadella et al.. 2013).
Recent studies in saltwater fish have examined the modifying effect of salinity. In chronic
exposure with larval topsmelt fish (A. ctffinis), Pb was consistently more toxic at lower salinity (14 ppt)
than at higher values (28 ppt) (Reynolds et al.. 2018). Free Pb2+ ion concentrations, the most bioavailable
form of Pb, were higher in the lower-salinity water, determined based on Pb speciation calculations in the
study. Lower-salinity water contains fewer cations, leading to decreased competition of free ionic Pb with
binding sites. Differential responses to salinity have also been reported in other studies in fish including
juvenile yellowfin seabream (Acanthopagnis latiis): the LC50 was significantly higher in fish acclimated
to 17 ppt salinity compared with fish acclimated to 0 ppt, 9 ppt, 25 ppt and 34 ppt salinity (Tsui et al..
2016).
11.4.2.4 Association with Sediments
Habitat type is a factor in the bioaccumulation of trace metals, as invertebrates closely associated
with benthic environments have greater contact with porewater and sediments, where metal
concentrations are higher than those in seawater. Several new studies published since the 2013 Pb ISA
reported differences in the biouptake of Pb associated with sediment characteristics. Bclzuncc-Scgarra et
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al. (2015) compared bioaccumulation in the benthic bivalve Tellina deltoidctlis with two sediment types
(silty, sandy) in the lab and deployed in the field. During the 31-day exposure period, Pb bioaccumulation
from sediments generally increased in a linear fashion with increasing sediment Pb concentration and was
greater in sandy sediments. For the silty sediments, there was more bioaccumulation in field-deployed
bivalves compared with bivalves in a parallel laboratory exposure, whereas the opposite was observed
with sandy sediments. Bioaccumulation in bivalves was attributed primarily to dietary exposure via
ingestion of particles due to the poor relationship between dissolved Pb in overlying waters (1 to
2.2 |ig Pb/L) and bioaccumulation. The authors noted that under laboratory exposure conditions, the
absence of processes occurring in the natural environment such as sediment resuspension, dilution of
surface sediments by deposition, and avoidance behaviors by organisms, likely lead to overestimation of
bioavailability. Battuello et al. (2018) quantified trace metals in two predaceous marine invertebrates
native to coastal waters of Italy: Enrydice spinigera (Isopoda), which burrows in sediments during the day
and rises to feed in the pelagic zone at night, and Flaccisagitta enflata (Chaetognatha), a zooplanktonic
species. Although the invertebrates have a similar feeding behavior and occupy the highest invertebrate
trophic level, Pb was an order of magnitude higher in E. spinigera (3.1 mg Pb/kg wet weight) compared
with F. enflata.
Fan et al. (2014) observed that the accumulation of Pb in polychaetes (marine annelid worms)
was significantly related to the total metal concentrations in sediment; however, metal concentrations in
polychaetes were less strongly correlated with metal concentrations in sediments if normalized for OC
concentration. The correlation improved when the metal concentrations in sediments were normalized for
Mn content, whereas normalization for Fe did not affect the correlation between Pb in sediment and Pb
accumulation in polychaetes. This suggested that Mn content in the sediment may be the driving factor
affecting bioaccumulation, while OM content in the sediment played little role in controlling the
bioaccumulation of Pb in polychaetes. Additionally, Pb accumulation in polychaetes was highly
positively correlated with its concentrations in FeMn oxides and organic fractions, and Pb
bioaccumulation in polychaetes was not related to its partitioning in different geochemical fractions.
11.4.2.5 Seasonality
Seasonal differences in Pb uptake and concentration in bivalves were noted in several European
field monitoring studies included in the 2013 Pb ISA (Carvalho et al.. 2011; Couture et al.. 2010; Pearce
and Mann. 2006). These differences could be due to seasonal changes in anthropogenic inputs or to
altered organism physiological condition in warmer versus colder months. Newer studies also reported
seasonal fluctuations in Pb uptake in saltwater invertebrates. Seasonal and spatial variation of trace metal
accumulation was observed in M. galloprovincialis mussels collected from sites around Port Phillip Bay,
Australia in the summer and winter (Shen et al.. 2020). In mussels collected from locations identified as
high risk for contamination, Pb body burden was higher in summer than in winter. In mussels collected
from less affected sites, there was no significant difference in Pb burden with season. This suggests that
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the increase in trace metals detected in mussels at more affected sites was due to greater anthropogenic
influence in summer. Metal bioaccumulation in red cherry shrimp (Neocctridinct denticulate!, now N.
davidi) sampled from a brackish wetland in Taiwan showed a seasonal variation in body residues, with
the highest accumulation of Pb in winter (Tang et al.. 2020). The saltwater shrimp could accumulate more
metal when wetlands shifted to a more heterotrophic system, as observed by the negative correlation
between net ecosystem production and Pb accumulation in shrimp. The highest ratios of Pb in shrimp to
waterborne Pb levels were found in winter (February), during the wetland"s highest season of
heterotrophy. Hernandez-Almaraz et al. (2016) measured heavy-metal content including Pb of white sea
urchins (Tripneustes depressus) and slate pencil sea urchins (Eucidaris thouarsii) collected in the
southwestern Gulf of California, Baja Sur California, Mexico in summer and winter and reported that Pb
concentrations were higher in E. thouarsii in the summer compared with the winter, likely due to diet.
11.4.2.6 Diet Composition
Few studies in saltwater biota have examined the role of diet composition on Pb uptake and
toxicity. Several studies in the 2013 Pb ISA reported tissue distribution patterns of Pb or assessed toxicity
to biota following dietary exposure (U.S. EPA. 2013). A study published since the 2013 Pb ISA
comparing the gut contents and Pb concentration of field-collected white sea urchins (T. depressus) and
slate pencil urchins (E. thouarsii) suggested different diets may influence Pb concentrations in these
organisms (Hernandez-Almaraz et al.. 2016). Specifically, Pb concentrations in the gonads of T.
depressus were below the detectable limit at all sites (<0.07 mg Pb/kg dry weight), while Pb
concentrations in the gonads of E. thouarsii ranged from 12.8 ±1.7 mg Pb/kg dry weight (mean ± SE) to
38.6 ± 4.2 mg Pb/kg dry weight. The diet for T. depressus varied with season and site and included both
brown and red macroalgae (mainly Sargassum, Gracilaria and Laurencia). The main food source for E.
thouarsii was red macroalgae, although they are considered a generalist omnivore that also fed on some
invertebrates, which was confirmed by higher S15N than T. depressus. Given Pb was only detected in E.
thouarsii, the authors suggested that these urchins might be exposed to Pb via macroalgae, specifically,
crustose macroalgae (Lithophyllum) or articulated coralline macroalgae (Amphiroa), as well as
invertebrates including mollusks, and/or barnacles.
In another dietary study Guo et al. (2013) examined whether the burned nassa sea snail
(Ncisscirius siquijorensis) showed differences in bioaccumulation patterns after being fed either Japanese
carpet shell clams (Ruditapes philippinarum), Asian green mussels (Pernct viridis), Fistulobalanus
albicostatus (barnacles) or Portuguese oysters (Crcissostreci angulata) for 8 weeks. The prey items were
collected from an intertidal zone in Xiamen, southeastern China. N. siquijorensis were sampled every
2 weeks and muscle and viscera metal concentrations, including Pb, were determined. In addition to the
body burden of metals in the snails, metal concentrations were also determined for the subcellular
fractions of the snails (heat-sensitive protein fraction, metallothionein-like protein fraction, MRG, cellular
debris and organelles). Pb concentrations differed between the four prey items (P. viridis:
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0.66 ±0.19 mg Pb/kg dry weight, mean + S.D., n = 8;R. philippinanim: 1.1 ±0.3 mg Pb/kg dry weight;
C. angidata: 2.4 ± 0.3 mg Pb/kg dry weight; F. cdbicostatus: 5.9 ± 1.1 mg Pb/kg dry weight). Subcellular
metal distribution in N. siquijorensis viscera and muscle at the beginning of dietary exposure was
concentrated in the cellular debris (44.3%). After exposure to four prey items over 8 weeks, the dominant
pool for Pb in the muscle was the cellular debris, while MRG became the dominant storage pool for
viscera across most prey items. Throughout feeding, MRG became a more important storage pool for Pb
relative to cellular debris. Pb was largely accumulated in the cellular debris and MRG for all prey items.
11.4.2.7 Lifestage
Additional studies on Pb effects in saltwater biota published since the 2013 Pb ISA provide
further evidence for variance in response to Pb at different lifestages. Embryo and juvenile lifestages are
commonly tested in bioassays due to their increased sensitivity to pollutant exposure. Many studies in
saltwater invertebrates discussed in the following sections continue to support findings in prior AQCDs
and the 2013 Pb ISA of differential toxicity with organism lifestage and increased sensitivity of larval or
other early lifestages compared with adults. In saltwater vertebrates, chronic toxicity bioassays with
topsmelt (A. affinis) at two lifestages (larvae and 2.5-month-old juveniles) lend further support to greater
sensitivity of earlier lifestages to Pb in saltwater fish (Reynolds et al.. 2018).
11.4.2.8 Historical Exposure
In the 2013 Pb ISA, the few studies that reported the development of tolerance to prolonged Pb
exposure were limited to freshwater invertebrates and fish: information was lacking for saltwater. A
recent study with the mangrove crab (Ucides cordcttus) collected from two locations in Brazil suggests
that a crab population inhabiting an historically polluted area may have developed mechanisms to cope
with elevated metals, resulting in differences in Pb accumulation compared with individuals from a
relatively pristine mangrove (Duarte et al.. 2020). After 28 days of laboratory exposure to low
concentration of Pb (10.6 |ig Pb/L), crabs collected from the protected site accumulated statistically
significantly more Pb in four of the six quantified tissues (gills, carapace, gonads, and muscle) and almost
double the total concentration of Pb compared with the crabs from the historically contaminated location.
The population from the protected site also took up more Pb in the biologically active form and exhibited
greater genotoxic effects (assessed by frequency of micronucleated cells and DNA strand breaks).
Furthermore, metallothionein induction in crabs from the historically contaminated location was more
than twice as high as that from the clean site.
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11.4.2.9 Species Sensitivity
As is the case for terrestrial and freshwater organisms, there are considerable differences in
response to Pb among saltwater biota. This information serves as the basis for the SSDs (Section 11.4.5)
for saltwater invertebrates and fish reported by (Church et al.. 2017). Both inter and intraspecific
differences in Pb uptake and bioaccumulation may occur in macroinvertebrates of the same functional
feeding group (U.S. EPA. 2013). For example, in the 2013 Pb ISA, data from 20 years of monitoring of
contaminant levels in filter-feeding mussels of the Mytilus genus and eastern oysters (C. virginica)
sampled along the U.S. coast, as part of the NOAA Mussel Watch program, indicate that Pb is on average
three times higher in mussels than in oysters (Kimbrough et al.. 2008). Wang et al. (2014b) compared
acute toxicity data (hazard toxicity ratios based on LC50 values; EC50 values for algal responses) for
temperate and tropical saltwater SSDs across five broad taxonomic groups (algae, crustaceans, fish,
mollusks, worms). Based on the hazardous concentration for 10% of the species (HC10) ratios, temperate
saltwater species are more sensitive to Pb than tropical saltwater biota. In the meta-analysis, algae were
the most sensitive taxa to Pb (HC10 = 29 |ig Pb/L, [95% CI 9.5, 86], n = 8) followed by fish
(HC10 = 166 ng Pb/L [95% CI 49], n = 10), crustaceans (HCK, = 428 pg Pb/L [95% CI, 263, 696], n = 22),
mollusks (HC10 = 1230 ^g Pb/L [95% CI 412, 3,660], n = 7), and worms (HC10 = 2,430 ^g Pb/L [95% CI
I,200, 4,610], n = 9).
II.4.2.10 Uptake and Bioaccumulation in Saltwater Plants and Algae
In the 1977 Pb AQCD, the cordgrass Spcirtinci alterniflora was found to reduce the quantity of Pb
in sediments by a small amount (U.S. EPA. 1977). Limited data on marine algae and saltwater plants
reviewed in the 1986 Pb AQCD, 2006 Pb AQCD, the 2013 Pb ISA and this appendix provide evidence
for species differences in Pb uptake and bioaccumulation rates.
One study examined element concentrations in pelagic Sargcissum that washed up along the coast
of the Yucatan peninsula in Mexico from the Caribbean (Rodriguez-Martinez et al.. 2020). Of 63 different
samples collected across eight sites from August 2018 to June 2019, only five samples had Pb levels at 2-
3 mg Pb/kg dry weight (as measured by X-ray fluorescence, which has a detection limit of 2 ppm). Other
metals such as As were detected in much higher amounts. Though Pb is not present in high amounts in
Sargcissum, the study showed that pelagic seaweed may be an avenue of transport across large distances
and contribute to Pb levels in coastal environments where it washes ashore.
An additional area of new research is the uptake of Pb by mangroves and the mechanisms that
may limit or confer tolerance. Mangrove swamps are coastal wetlands found in tropical and subtropical
regions. They are characterized by halophytic woody plants growing in brackish to saline tidal waters.
One greenhouse experiment aimed to investigate the possible function of root lignification/suberization
on Pb uptake and tolerance in two pacific mangrove species with different degrees of root lignification
and suberization: holly mangrove (Acanthus ilicifolius) and red mangrove (Rhizophora stylosa) (Cheng et
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al.. 2015). Plants were grown in pots with three nominal Pb treatments applied to the sediment—low
(250 mg Pb/kg), medium (500 mg Pb/kg) and high (1,000 mg Pb/kg)—and one control with no Pb; Pb
exposure was a period of 3 months. In the species with little lignification and suberization, A. ilicifolius,
biomass yield decreased significantly as plants were exposed to increasing concentrations of Pb; about 20,
35 and 50% reductions were observed in low, medium, and high Pb treatments when compared with the
respective controls. R. stylosct, however, was not affected by low and medium Pb exposure. A significant
decrease in relative Pb was observed within the outer cortex cell layers, indicating that lignified/suberized
exodermis acts as a barrier to the movement of Pb. A further study with six pacific mangrove species
subjected to different levels of a metal mixture (Pb with Zn and Cu) corroborates these findings and
suggests that mangrove species, which possess more extensive lignification and suberization within their
root exodermis, exhibit higher tolerance for heavy metals (Cheng et al., 2014).
The U.S. EPA Framework for Metals Risk Assessment states that the latest scientific data on
bioaccumulation do not currently support the use of BCFs and BAFs when applied as generic threshold
criteria for the hazard potential of metals (U.S. EPA, 2007); however, such metrics are useful to provide
information about the amount of uptake of metals into plants, compartmentalization into different plant
tissues, and differences between species. In a field study conducted in four marine and four inland
wetlands in Sicily with differing levels of anthropogenic impacts, Pb concentrations were quantified in
soil, water, and plant tissues of two Mediterranean seagrasses, Posidonict Oceania and Cymodocea
nodosa, and five freshwater species were quantified (Bonanno et al., 2017). Sediment Pb levels ranged
from 2.56 ± 0.33 mg Pb/kg at the lowest impacted site to 11.5 ± 1.57 mg Pb/kg at the most impacted site
for the marine sites and 1.05 ± 0.21 to 17.2 ± 4.58 mg Pb/kg for the freshwater sites. BCFs (C root/Csediment)
were higher for the two marine seagrasses than those for any of the freshwater species, more than twice as
high as the values for the highest freshwater species (0.71 and 0.84 for/1. Oceania and C. nodosa,
respectively, compared with 0.03-0.30 for the freshwater species). For both marine species, Pb was
concentrated in root tissue, but translocation factors into different tissues differed between species. An
additional study Bonanno et al. (2020) examining the seagrass C. nodosa and marine green algae Ulva
lactuca showed that both species are comparable in their ability to sequester high levels of trace elements
including Pb.
11.4.2.11 Uptake and Bioaccumulation in Saltwater Invertebrates
At the time of the 1977 AQCD, it was understood that shellfish concentrate Pb in their tissues and
shells (U.S. EPA, 1977). Uptake and subsequent bioaccumulation of Pb in marine invertebrates varies
greatly between species and across taxa as previously characterized in the 2006 AQCD (U.S. EPA, 2006)
and the 2013 Pb ISA (U.S. EPA, 2013). In the case of invertebrates, Pb can be bioaccumulated from
multiple sources, including the water column, sediment, porewater and dietary exposures, and factors
such as the proportion of bioavailable Pb, lifestage, age and metabolism can alter the accumulation
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rate. Since the 2013 Pb ISA additional information on uptake rates, Pb sequestration patterns and Pb
accumulation from aqueous and dietary exposures has been published for saltwater invertebrates.
As reported in studies in previous reviews, major sites for Pb accumulation following aqueous
exposure include the gill and digestive gland or hepatopancreas; current studies continue to support these
findings. In pacific oyster (C. gigas) exposed nominally to 5 |ig Pb/L for 9 days, Pb concentration in the
gill and digestive glands were 19- and 24-fold higher, respectively, than Pb measured at the beginning of
the experiment (Meng et al.. 2018). Following 28-day exposure to a low concentration of Pb
(10.6 |ig Pb/L), the highest concentration of Pb was accumulated in the gill, followed by the carapace in
the mangrove crab (U. cordcttus) (Duarte et al.. 2020). The crabs sequestered Pb in detoxified forms, with
differences in Pb accumulation and storage observed in two distinct populations (crabs collected from a
protected mangrove area and those collected from a historically contaminated site).
Adult female Atlantic Horseshoe crabs (Limulus polyphemiis) collected from several different
beaches in Long Island, NY, had higher Pb concentration in gills than legs or eggs; Pb in leg tissue was
significantly and positively correlated with egg Pb burden, suggesting maternal transfer of the internalized
metal to eggs (Bakker et al.. 2017b). Pb quantified in field-collected horseshoe crab embryos (range 0.05-
0.43 mg Pb/kg dry weight) and developing larvae (range 0.07-0.59 mg Pb/kg dry weight) was compared
with Pb concentration in eggs, sediment, porewater and overlaying water (Bakker et al.. 2017a). Although
Pb measured in environmental media varied between sites, the concentration of Pb significantly increased
from egg to embryo at four out of five sampling locations, indicating uptake of Pb from the surrounding
substrate following hatching since the embryonic lifestage develops in the sediments. There was no
significant change in Pb concentration when comparing embryos to larvae; however, the authors noted
that it is possible some trace metals are lost at the larval stage during molting.
Embryos of the sea urchin S. piirpiiratus exposed to an analytically verified Pb concentration of
55 |Lig Pb/L during 96-hour embryo toxicity assays showed significant Pb accumulation after 12 hours
through 96 hours of development, with a peak at 84 hours (Tellis et al.. 2014). Pb disrupted Ca uptake
during initial development stages, especially during gastrulation, and there was a corresponding increase
in Ca2+ATPase activity in the embryos; however, Ca levels in Pb-exposed embryos returned to control
amounts by 72 hours.
A few dietary exposure studies in marine invertebrates have been conducted since the 2013 Pb
ISA. In sea hare (Aplysia ccdifornica) exposed to Pb solely through diet (green seaweed U. lactiica
previously exposed to an analytically verified concentration of either 10 |ig Pb/L or 100 |ig Pb/L for
48 hours), the Pb accumulation pattern in the mollusk was greatest in the hepatopancreas followed by the
gill and crop (Jarvis et al.. 2015). In sea cucumbers (Apostichopus jctponiciis) fed a Pb-supplemented diet
(100, 500, or 1,000 mg Pb/kg dry weight) for 30 days, the profile of tissue Pb accumulation was body
wall>intestine>respiratory tree (Wang et al.. 2015). The bioavailability of Pb from food and subsequent
trophic transfer is affected by how Pb is stored in the prey. In a feeding study, the common prawn
Pcdaemon serratus was fed for 28 days with either tissues from the mussel (M. galloprovincialis) exposed
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to 100 |ig Pb/L for 48 hours or tissues from the field-collected clam Dosinia exolete, wherein Pb is stored
primarily in nonbioavailable MRG in the kidney (Sanchez-Marin and Beiras. 2017). Although the Pb
concentration in both food items was similar (15 and 17 mg Pb/kg wet weight, respectively), Pb
accumulation in prawns was 10x higher when fed tissue from the mussels, in which Pb was in a more
soluble subcellular faction, compared with the prawns consuming D. exolete, in which Pb was in a less
bioavailable form.
Pb uptake is influenced by feeding strategy. In the filter-feeding bivalve Andctrct trapezia, uptake,
and bioaccumulation from Pb-spiked sediments (analytically verified concentration of 100 and
300 mg Pb/kg) to the gill and mantle, hemolymph and hepatopancreas were quantified on days 0, 14, 28,
42 and 56 of a 56-day exposure (Taylor and Maher. 2012). At the end of the experiment, total Pb
concentration in the mollusk was 1 mg Pb/kg at the low concentration and 12 mg Pb/kg at the high
concentration. In the highest Pb treatment, an increase in Pb in hemolymph was observed from day 42 to
day 56, resulting in a doubling of Pb tissue concentration. The authors speculated this could be related to
greater availability of dissolved Pb in porewater over time due to oxidation of the sediments. Generally,
the order of tissue accumulation was hemolymph > gill and mantle > hepatopancreas over the 56-day
exposure. In contrast, the deposit-feeding bivalve Tellina deltoidalis exhibited a distinct pattern of Pb
uptake under similar experimental conditions and exposure to spiked sediments (28-day exposure to
analytically verified concentrations of 100 and 300 mg Pb/kg) (Taylor and Maher. 2014). Individuals in
the 100 mg/kb Pb-spiked sediment rapidly accumulated Pb early in the exposure period (day 3) followed
by continued uptake over the remainder of the experiment, to reach a final tissue concentration
(96 mg Pb/kg) equal to that of the spiked sediment. In the 300 mg Pb/kg microcosm, the bivalves seemed
to exhibit a pattern of uptake and loss over the 28-day period, with the highest Pb concentration at day 21
and a final total Pb concentration of 430 mg Pb/kg.
Aquatic invertebrate strategies for detoxifying Pb reviewed in the 2006 Pb AQCD and 2013 Pb
ISA included sequestration of Pb in lysosomal-vacuolar systems, excretion of Pb by some organisms and
deposition of Pb to molted exoskeleton. Pb can be stored in two forms: biologically detoxified metal
(which includes MRG) and biologically available metal. Following the biouptake experiments described
above, subcellular partitioning of Pb was determined in the bivalves (Taylor and Maher. 2014. 2012). Of
the recovered Pb in A. trapezia tissues, Pb was associated to the greatest extent with the biologically
detoxified metal fraction (ranging from 66% to 69% in the gill and mantle to 56% in the hepatopancreas),
distributed fairly evenly between the metallothionein-like proteins and MRG fractions (Taylor and Maher.
2012). In T. deltoidalis, Pb was also primarily found in the biologically detoxified metal fraction
(approximately 70%), with 74% of the total detoxified Pb converted to MRG and the remainder in the
metallothionein-like protein fraction (Taylor and Maher. 2014).
In a 96-hour exposure to analytically verified concentrations of Pb (0-1,800 |ig Pb/L),
intracellular partitioning data in adult clams Venernpis decussata showed that most Pb accumulated in the
insoluble fraction (>80%), a form not readily bioavailable to consumers at higher trophic levels (Freitas et
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al.. 2014). Total Pb in clams increased with increasing water concentration up to 230 |ig Pb/L, then
decreased at higher concentrations. The clams bioconcentrated Pb in the soluble fraction more efficiently
at low water concentrations (BCF > 26) compared with higher concentrations (>450 |ig Pb/L; BCF < 16).
11.4.2.12 Uptake and Bioaccumulation in Saltwater Vertebrates
Studies reviewed in prior AQCDs and ISAs report Pb accumulation in tissues sampled from
seabirds, saltwater fish, and marine mammals (U.S. EPA, 2013, 2006, 1977); however, there are fewer
biouptake studies of Pb in saltwater than in freshwater. Because marine fish drink seawater to maintain
osmotic homeostasis, Pb can be taken up from the water column via both the gills and intestine (Lee et al.,
2019; Wang and Rainbow, 2008). In the 2013 Pb ISA, storage of Pb in metal granules was reported as a
detoxifying mechanism in mummichogs (Fundulus heteroclitus). Fish at more polluted sites stored a
higher amount of Pb in MRG as compared with other detoxifying cellular components such as heat-stable
proteins, heat-denaturable proteins and organelles (Goto and Wallace, 2010). Since the 2013 Pb ISA,
additional studies have further elucidated the role of subcellular fractions in metal detoxification in
saltwater fish. Metal binding to subcellular fractions in the livers of wild-caught yelloweye rockfish
(Sebcistes ruberrimus) collected from the southeast coast of Alaska was assessed to gain a better
understanding of the degree to which this long-lived endangered fish species can detoxify nonessential
metals including Pb (Barst et al„ 2018). Combining data from the rockfish, Pb was detected to a greater
extent in the detoxified compartment (46%); however, detoxification was incomplete given that Pb was
also present in metal-sensitive fractions (a total of 35%, divided between heat denatured proteins [12.2%],
mitochondria [11.4%], microsomes and lysosomes [10.8%]). Metals associated with sensitive subcellular
fractions indicate a risk of disruption to cellular processes; however, the concentrations of Pb in rockfish
were low compared with other detected metals. These patterns were consistent with results from
subcellular partitioning in livers of yellow eels native to North America (Anguilla rostrcitci) and Europe
(Anguillct anguilla) (Rosabal et al., 2015). In both eel species, the granule-like detoxification fraction
showed the strongest increase in Pb concentrations among all subcellular fractions, with the metal-
sensitive mitochondrial fraction representing a significant binding compartment for Pb.
A novel study exploring the use of fish eyes as an organ for monitoring Pb exposure compared Pb
concentration in mullet (Liza aurata) eyes, water column and sediment in a metal-contaminated location
and reference area within the same estuary (Pereiraet al„ 2013). Eyes from individuals collected from the
contaminated site (0.81 |ig Pb/L water column, 417 mg Pb/kg sediment) had significantly higher Pb
accumulation (10x) than the less affected site (0.032 |ig Pb/L water column, 61 mg Pb/kg sediment),
suggesting the eye is a target organ for Pb. It is not known if the accumulation of metals in the eye is from
direct contact with water or redistribution of Pb taken up by the fish via other routes of exposure.
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Studies that considered uptake of Pb in saltwater birds and mammals are limited to surveys of
field-collected individuals that reported Pb concentration in tissue or trace-element patterns of tissue
distribution.
11.4.2.13 Uptake and Bioaccumulation Through Marine Food Web
Trophic transfer of Pb in marine food webs was found to be negligible in the 2006 Pb AQCD
(U.S. EPA, 2006) and the 2013 Pb ISA (U.S. EPA, 2013). In many studies reported in previous
assessments and those reviewed here, Pb was found to decrease with increasing trophic levels, although
some studies found evidence of bioaccumulation. Whether Pb is biodiluted or bioaccumulated in marine
food webs depends on the sediment and porewater Pb, the type of marine ecosystem, the organisms
examined, and other contaminants. In a review published in 2013, Cardwell et al. (2013) compiled
laboratory and field studies to examine the transfer of Pb and other heavy metals through marine food
webs. In most of the field studies reviewed, no evidence was found for biomagnification of Pb across
trophic levels. Specifically, nine studies examined trophic transfer of heavy metals through marine food
webs in the field. Eight of these studies found no evidence of biomagnification of Pb, and one did not
examine Pb or did not present data on Pb. More recent studies are presented below.
Biodilution of Pb in marine food webs was supported by an environmental gradient study on a
green sea turtle food web in San Diego Bay, California, U.S. (Komoroske et al., 2012). Green sea turtles
(Chelonia mydcts) largely forage on eelgrass (Zostera marina) and invertebrates, and exposure to heavy
metals occurs primarily through foraging, as these organisms breathe air and do not feed during
migration. At each of eight eelgrass sites, sediment samples, eelgrass, red algae (Gracilaria spp.), green
algae (Ulva spp.), soft-bodied invertebrates (i.e., Zoobotryon spp.), sponges, and green sea turtle carapace
tissues were collected and analyzed for trace metals. Mean Pb concentrations in sediments and organisms
varied across season and site in San Diego Bay. Pb did not bioaccumulate in eelgrass or algae: Pb in the
sediment was significantly higher than Pb in eelgrass and red algae, but not higher in green algae.
Biodilution of Pb was also reported across six intertidal sites in New England (four in the Gulf of
Maine and two in Narragansett Bay, Rhode Island) spanning a gradient of watershed land use and
urbanization (Chen et al., 2016a). Sediments, invertebrates, and benthic and pelagic fish were sampled
and analyzed for heavy metals. Trophic position was characterized using stable-isotope analysis on biotic
tissues. Specifically, S13C is correlated with the relative proportion of pelagic diet sources, while S15N is
related to trophic position. Invertebrate and fish samples were categorized into five taxonomic groups, as
the same species were not collected at all sites. Biota-sediment accumulation factors (BSAFs) were
calculated for each taxonomic group (amphipod, crab, Fundulus, mussel, and shrimp) as the metal
concentration in the organisms divided by the metal concentration in the sediment. Positive logio BSAF
values indicate bioaccumulation, while negative values indicate biodilution. Pb concentrations in the
sediment across six sites in the Gulf of Maine and Narragansett Bay ranged from 4.7 mg Pb/kg to
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79.6 mg Pb/kg, and these concentrations increased linearly with the percent of total OC. All logio BSAF
values were negative for Pb, indicating that organisms in higher trophic levels contained less Pb than
organisms occupying lower trophic levels. Pb concentration across five taxonomic groups (mussel,
shrimp, crab, Fundulus, and amphipod) showed considerable variation across taxa and sites.
Simultaneously extracted metal-AVS in the sediment were marginally predictive of biota Pb content,
while trophic level and pelagic feeding were not predictive of biota Pb.
Trophic level positions of a marine invertebrate community and body Pb concentrations of a
marine invertebrate community were not correlated in the Bay of Fundy, Nova Scotia, Canada suggesting
Pb does not bioaccumulate in this system (English et al.. 2015). The invertebrate community included
barnacles (Balanus balanus), worms (Cerebratulus lacteus, Clymenella torqucitci, Glycera dibrcinchicite,
Hediste diversicolor), amphipods (Corophium volutator, Gammctrus oceanicus), clams (Ensis directus,
Mva arenaria, Mctcoma balthica), snails (Ilyanassa obsolete,i, Littorina littorea), mussels (Mytilas ediilis),
and crabs (Pa gurus pubescens). Stable isotopes were used to characterize the relative trophic position as
organisms in higher trophic levels contain higher levels of S15N, while S13C is often associated with lower
trophic levels. Although the species sampled likely did not belong to the same food web, they occupy
similar trophic levels in different food webs and are all important prey items for species in higher trophic
levels. In this study, S15N was negatively correlated with S13C for most species. Pearson correlation
coefficients were calculated between stable-isotope and trace-element content for each species to test for
bioaccumulation or biodilution through the food web. Pb concentration varied among invertebrate species
in the community; however, no single species had higher Pb concentrations than the others. Pb ranged
from 0.07 ± 0.01 mg Pb/kg (mean ± S.D.) in Glycera dibranchiate (Polychaeta) to 1.25 ± 1.40 mg Pb/kg
in I. obsoleta (Gastropoda). There were no significant correlations between trophic level position (S15N or
S13C) and logio Pb concentration, suggesting Pb does not show considerable bioaccumulation in the food
web.
In another example, trophic level position determined using stable isotopes of white sea urchins
(T. depressus), slate pencil sea urchins (E. thonarsii), and nine types of macroalgae food sources in four
Sargassum beds in the southwestern Gulf of California in Baja California Sur, Mexico were not correlated
(Hernandez-Almaraz et al.. 2016). Out of the macroalgae and two sea urchins studied, E. thoiiarsii had
the highest Pb concentrations (ranging from 12.8 ± 1.7 mg Pb/kg dry weight [mean ± SE] to
38.6 ± 4.2 mg Pb/kg dry weight) and stable-isotope content.
Pb accumulation in a tropical estuarine lagoon in Mexico decreased with increasing trophic level
(Mendoza-Carranza et al.. 2016). Sediment Pb concentration was 20.86 ± 5.80 mg Pb/kg (mean ± S.D.),
and the suspended load Pb concentration was 16.59 ± 2.79 mg Pb/kg in the San Pedrito Lagoon, which is
impacted by wastewater discharge and petroleum extraction. Pb concentration was only above the limit of
detection in two plant species, spider lily (Hymenocallis littoralis) and mangrove fern (Acrostichum
aiireum), and fish samples (2.9 mg Pb/kg) from the lagoon. In general, BCFs of Pb were low, and BCFs
were higher in plants than in fish, suggesting trophic dilution.
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Bioaccumulation, but not biomagnification, was found in a semiarid coastal lagoon in Sonora,
Mexico along the Gulf of California (Jara-Marini et al.. 2020). The community consists of primary
producers (phytoplankton, algae, mangrove), primary consumers (zooplankton, barnacles, oysters, clams,
snails, shrimp, crab, snapper, and juveniles of flathead mullet), secondary consumers (adults of flathead
mullets, crab, snapper, mojarra, and grunt), and tertiary consumers (night herons, great blue herons,
magnificent frigate, and cormorants). BMF was corrected for the trophic position, and the trophic BMF
(TBMF) was estimated from the antilogarithm of the slopes of the linear correlation between the trophic
level and the metal concentration. BMF and TBMF values above 1.0 indicate that a metal is being
transferred upward through the trophic levels, while values below 1.0 indicate biodilution along the food
web. Pb values in suspended particulate matter and sediment varied between seasons, ranging from
0.70 mg Pb/kg in autumn to 1.03 mg Pb/kg in winter. Pb concentrations among primary producers (range:
0.63 to 1.03 mg Pb/kg), secondary consumers (range: 0.80 to 1.53 mg Pb/kg), and most tertiary
consumers did not vary seasonally. Only two tertiary consumers, neotropic cormorant (Phalacrocorax
brasilamis) and magnificent frigatebird (Fregata magnificens), showed the highest Pb concentrations in
the summer. Pb only showed a positive relationship between log-transformed Pb concentrations and
trophic level in the summer. Pb concentrations generally decreased through the food web, depending on
the season. The BMF ranged from 0.50 to 3.57 for Pb across organisms, and TBMF ranged from 1.02 to
1.15. Although TBMF values were above 1.0, biomagnification was unlikely because the relationship
between trophic level and Pb concentration was only significantly positively correlated in the summer.
In addition to evidence from field studies, laboratory findings also suggest a decrease in the
concentration of Pb with trophic transfer. In an 8-week feeding study, TTFs were calculated for sea snail
(Nassarius siqnijorensis) fed either venerid clams (Rnditapes philippinariim), mussels (Pernct viridis),
barnacles (Fistulobalanus cdbicostatiis), or oysters (Crcissostreci. angiilata) collected from an intertidal
zone in Xiamen, southeastern China (Guo et al.. 2013). The net trophic transfer factor, which is the ratio
of net accumulated metal concentrations over the experiment to metal concentrations in the prey was well
below 1 for barnacles and 0 for oysters, clams, and mussels, suggesting biodilution in this system.
Although not tested statistically, the variation in trophic transfer factors across prey species demonstrated
prey-specific bioavailability.
Although most observational studies suggest biodilution of Pb occurs through marine food webs,
Pb was found to bioaccumulate in mummichog (Fundulus. heteroclitns) in the Goose Pond estuary in
Brooksville, Maine (Broadlev et al.. 2013). The Goose Pond estuary was impacted by the former Callahan
Mine, which is one of the few documented open-pit hard-rock mining sites in an intertidal zone. The
sediment concentration of Pb was above the probable effects level in some of the sample sites. BAFs
ranged from 3.3 to 3.65 across Goose Pond sites and from 2.62 to 3.77 at the reference sites. The
reference site values bioaccumulation factors were conservative as they included water and tissue Pb
concentrations below the instrument detection limit. The sediment-to-F. heteroclitns and water-to-/''.
heteroclitns ratios were high for Pb. The ratio of metal enrichment to background levels (concentrations at
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the Goose Pond site adjacent to the tailings pile / the mean concentrations at a reference site) were 34.2
for sediment, 32.3 for water, and 45.6 for F. heteroclitiis.
Environmental gradient field studies outside of North America provide additional evidence to
support the biodilution of Pb in marine food webs. For example, trophic transfer of Pb was low in a
seagrass food web in an estuarine lake in Australia, as the trophic level was negatively correlated with Pb
concentration (Schneider et al.. 2018). In another example, Pb concentrations decreased with increasing
trophic level in a Mediterranean coastal lagoon (Vizzini et al.. 2013) and in a small pelagic fish marine
food web along a Mediterranean coast (Chouvclon et al.. 2019). Finally, Pb accumulation was higher in
invertebrates compared with higher trophic level species (fish) in an aquatic food web in Liaodong Bay,
China (Radomvski et al.. 2018).
In summary, studies published since the 2013 Pb ISA support findings in the ISA that Pb
generally decreased with increasing trophic level in coastal and marine food webs, although some studies
found evidence of bioaccumulation.
11.4.3 Environmental Concentrations of Pb in Saltwater Biota in the United
States at Different Locations and Over Time
Studies of aquatic bivalves in coastal ecosystems can be used to reconstruct historical records of
Pb concentrations. The NOAA Mussel Watch program has monitored pollutant trends since 1986 via
periodic sampling of bivalve tissue (Mytilus species and C. virginica oysters) and sediment along the U.S.
coastline (Kimbrough et al.. 2008). In general, the highest concentrations of Pb are in bivalves in the
vicinity of urban and industrial areas, and Pb is, on average, three times higher in coastal mussels than in
oysters. Metal concentrations in Mytilus californiamis were sampled at long-term biomonitoring sites off
the coast of California from 1977 to 2010 (specific years vary by site) as part of the National Mussel
Watch (NSW) (n = 35 sites) and California State Mussel Watch (CSMW) (n = 21 sites) (Melwani et al..
2014). Decreasing trends were observed at 11 NMW sites and 8 of the CSMW sites; no significant trends
were found at the remaining sites. These observations show that Pb inputs to coastal aquatic ecosystems
from runoff have decreased significantly, especially at sites off the coast of southern California near large
municipal wastewater treatment facilities.
Quantification of chemical variation in relative presence of Pb and of other elements taken up and
deposited in shells of marine organisms (sclerochronology) provides a temporal record of Pb deposition
inputs to coastal environments. In a 2005 study of Mercenaria shells collected off the coast of Cape
Lookout, North Carolina in 1980, 1982, 2002 and 2003, annual average Pb/Ca ratios were estimated from
1949 to 2002 using concentration measurements milled between the mollusk shell growth lines, which
provide corresponding chronological measurements (Gillikin et al.. 2005). Although high variability
between samples was observed, overall Pb/Ca ratios in shells peaked near 1980 and decreased until the
conclusion of the sampling in 2003. This study provides an indicator that reductions in Pb pollution
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resulted in decreased Pb inputs to aquatic ecosystems through runoff on the east coast. Elemental analysis
of shell carbonate of the long-lived bivalve Arctica islctndicct collected off the coast of Virginia revealed a
pattern of continuous increase in Pb concentration after 1910, reaching a peak in 1979 and declining after
that date to pre-1930 values after 2000 (Krausc-Nchring et al.. 2012). The elevated shell Pb corresponded
to the period of peak leaded gasoline use in the United States, with Pb deposition to the offshore site
including atmospheric transport by easterly winds. Cariou et al. (2017) synthesized data from 15 studies
from different geographic locations that quantified Pb in marine bivalve shells. They found that shell
concentration had a strong relationship with the environmental level of local contamination; values in the
shells, which ranged from 0.08 mg Pb/kg to 2 mg Pb/kg, were associated with environments with distant
Pb sources including atmospheric deposition.
In addition to bivalve tissue and shell, heavy metals in horseshoe crab (L. polyphemns) eggs
collected from breeding grounds on beaches along Delaware Bay provide some historical data for trend
analysis. Horseshoe crab eggs collected in 1993, 1994, 1995, 1999, 2000, and 2012 showed a decline in
Pb overtime in a comparison of compiled data from the earlier surveys (1993, 1994, 1995)
(x = 0.289 ± 0.068 mg Pb/kg) and to the data from 1999 to 2000 (x = 0.0353 ± 0.00496 mg Pb/kg)
(Burger and Tsipoura. 2014). Some of the individual resampled sites showed a clear temporal decrease in
Pb from 1993 to 2012, while at other locations, the temporal Pb concentration trend was more variable.
A study of migratory shorebird species in Delaware Bay compared feather Pb concentrations
from the 1990s with samples from 2011 and 2012 (Burger et al.. 2015). The decline of shorebirds
migrating through Delaware Bay over the study period has been widely attributed to the reduced size of
horseshoe crab populations, whose eggs the migratory birds feed on. Declining populations have been
observed elsewhere in the shorebirds" ranges, and the authors investigated heavy metals as a driver of
those declines. Across the time period studied, Pb concentrations increased in red knots (Calidrus
canutus), decreased in semipalmated sandpipers (Calidrus piisilla), and did not change significantly in
sanderlings (Calidris alba) (Burger et al.. 2015). The authors noted that Pb concentrations observed in
this study were below the known adverse effect risk levels for similar species.
In a decade-long biomonitoring study of metals in the muscle tissue of dolphinfish (Coryphaena
hippiinis) in the southern Gulf of California from 2006 to 2015, Gil-Manrique et al. (2022) found no
temporal trend in Pb concentrations. However, a negative correlation was identified between sea surface
temperature and Pb concentrations in dolphinfish during the decade-long biomonitoring study. Summary
statistics of dolphinfish sampled in Gil-Manrique et al. (2022) are included in Table 11-1.
In long-term biomonitoring studies of saltwater ecosystems, there is some evidence of declining
Pb concentrations, particularly in studies which began sampling before the 1990s. However, other studies
document mixed results, with some observations of insignificant change or even increases in Pb
concentrations.
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11.4.4
Effects of Pb in Saltwater Systems
Saltwater taxa included in this section are broadly grouped into vegetation, microbes,
invertebrates, and vertebrates. The biological effects of Pb in the 2013 Pb ISA and in this appendix are
generally presented from suborganismal responses (i.e., enzymatic activities, changes in blood
parameters) to endpoints relevant to the population level and higher (growth, reproduction, and survival)
up to effects on ecological communities and ecosystems. Exposure-response studies that report
toxicological dose descriptors (e.g., LC50, EC50, LOAEL) for effects on growth, reproduction or survival
endpoints are reported in Section 11.4.5.
11.4.4.1 Effects on Saltwater M icrobes
Microbial communities in saltwater ecosystems were not reviewed in detail in the 2006 Pb
AQCD (U.S. EPA, 2006) or the 2013 Pb ISA (U.S. EPA, 2013). More recent experimental and
observational studies reviewed here examine the relationship between Pb concentration in the sediment
and saltwater and the effects on marine microbial communities. Pb was largely negatively or not
associated with microbial community structure and abundance, although a few studies found positive
associations between sediment Pb concentrations and microbial abundance.
Pb negatively affected microbial diversity and structure in rhizosediments of sea rush (Juncus
maritimus) and the common reed (P. australis) collected from the Lima estuary, Portugal and exposed to
a Pb gradient (Mucha et al., 2013). Juncus maritimus rhizosediment exhibited lower OTU number,
diversity, evenness, and dominance in Pb-exposed sediments relative to the control. Similarly, P.
australis rhizosediments exposed to Pb exhibited lower OTU number, diversity, and evenness than the
control, whereas dominance was unaffected by Pb exposure. Both rhizosediment microbial community
structures under 218 mg Pb/kg and 2180 mg Pb/kg were dissimilar to the controls and to one another.
Sediment Pb concentration was correlated with bacterial richness and evenness along a gradient
of metal pollution in estuaries on the southeast Australian coast (Sun et al., 2012). The relationships
between sediment heavy-metal content (Pb, Cr, Cu, Fe, Mn, Ni, Pb, and Zn), organic contaminants
(polycyclic aromatic hydrocarbons), physicochemical variables (silt content and OM), water column
environmental parameters (temperature, pH, and salinity) and bacterial community structure were
explored using Automated Ribosomal Intergenic Spacer Analysis profiles across six sites with different
degrees of anthropogenic disturbance. High collinearity existed between silt content and Cr, Ni, Zn, and
Pb; therefore, only latitude, salinity, temperature, pH, %silt, Cu, and Zn were included in the analysis of
bacterial community structure. Silt, which was highly correlated with Pb concentration, was the main
driver of bacterial community structure, followed by temperature. Pb explained the highest relative
proportion of variance in bacterial diversity (16.1% explained by Pb followed by 14.5% by Cu and 7.5%
by silt), and bacterial diversity decreased with increasing Pb and Cu sediment concentration.
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The relative abundances of certain bacterial groups were negatively correlated with the Pb
sediment concentration of mangroves in southern China (Meng et al.. 2021). Sediment Pb concentrations
ranged from 0.142 ± 0.094 mg Pb/kg to 3.257 ± 0.094 mg Pb/kg (mean ± S.D.) across seven sites, and the
mean Pb concentrations in surface sediments (0-5 cm) were higher than those in deep sediments (25-
30 cm). Pb was significantly correlated with Zn sediment concentration. The abundance of the genus
Fusibcicter was negatively correlated to Pb, Zn, Cu, Co, Ni, Cd, and Ag with statistical significance, while
Svntrophorhabdiis was positively correlated with Pb. Among the >200 genera and functional genes
involved in heavy-metal transportation, most bacteria associated with Pb elimination and transport
demonstrated lower abundances compared with other genera.
Although some studies report negative correlations between Pb sediment concentration and
bacterial community structure, other studies found no such relationship. For example, Pb sediment
concentration was not correlated with bacterial community structure in the Jiaozhou Bay, China (Yaoet
al.. 2017). Sediment samples were collected from inside Jiaozhou Bay and a site outside of the bay to
achieve an environmental gradient of water quality.. The concentration of Pb varied among the three sites
(mean ± S.D. site Shilaoren Beach: 19.09 ± 1.86 mg Pb/kg, site Haibohe estuaries: 38.65 ± 9.26 mg Pb/kg
and site Licunhe estuaries: 72.87 ± 17.56 mg Pb/kg). The concentrations of Co, Zn, Hg, As, and Se
explained more of the variation in bacterial community composition in this system, and Pb concentration
was not one of the top three strongest predictors of bacterial diversity at any site.
In eastern Guangdong, China, marine microbial community diversity was not correlated with Pb
content; however, Pb was significantly correlated with the abundance of a few dominant taxa (Zhuang et
al.. 2019). Pb in the sediment from the Shantou coastline ranged from 4.9 mg Pb/kg to 95.7 mg Pb/kg,
with a mean value of 37.04 mg Pb/kg. The only significant correlations between bacterial diversity and
abundance and environmental variables were total OC and Cr, and the correlation between Pb and
microbial abundance and diversity was not significant. Although bacterial diversity and abundance were
not correlated with Pb, the metal was significantly negatively correlated with the abundance of
Nitrospirae and positively correlated with candidate phylum OD1. Additionally, sediment Pb
concentration was significantly negatively correlated with a few dominant classes, including Epsilon-
proteobacteria, Nitrospira, and Sva0725. Given there was a significant positive correlation between OD1
and all metals and a negative relationship between Nitrospirae and all metals and Pb was highly correlated
with other heavy metals (As, Hg, Cu, Zn, and total OC), it is difficult to disentangle the sole effects of Pb
on marine microbial communities.
Finally, Pb concentration in the water column did not affect bacterioplankton community
composition in the Toulon Bay, France (Coclet et al.. 2019). Mn, DOC, salinity, Cu, and Cd explained the
most variation in the bacterioplankton community composition (range in variance: 1.01%—1.22%), while
Pb concentration only explained 0.51% of the variance in bacterioplankton community composition.
Rhodobacteraceae, SARI 1 (Alphaproteobacteria), Balneola (Bacteroidetes), and Synechococcus
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(Cyanobacteria) were negatively correlated with either Cd, Cu, Pb or Zn, while Candidatus aqnilina
(Actinobacteria) was positively correlated with Pb.
In summary, several experimental and observational studies since the 2013 Pb ISA (U.S. EPA.
2013) reported negative relationships between sediment or saltwater Pb concentration and microbial
abundance and diversity (Meng et al.. 2021; Mucha et al.. 2013). while other studies found no relationship
(Coclet et al.. 2019; Zhuang et al.. 2019; Yao et al.. 2017).
11.4.4.2 Effects on Saltwater Plants and Algae
In the 2013 Pb ISA, evidence was inadequate to infer a causal relationship between Pb exposure
and endpoints relevant to saltwater plants and algae (growth, survival, physiological stress) (Table 11-6).
Key studies in the 2013 Pb ISA included a 72-hour EC50 for growth inhibition reported in the marine
algae Chaetoceros sp. at 105 |ig Pb/L (Debelius et al.. 2009). A study with the microalga Tetrctselmis
snecica reported a statistically significant decrease in growth rate, total dry biomass and final cell
concentration between control cultures and algae cultured in 20 |ig Pb/L (Soto-Jimenez et al.. 2011). Few
data are available in prior Pb reviews for saltwater plant and algal species. Effects in plants, in general,
are observed at concentrations of Pb that greatly exceed concentrations of this metal typically measured in
soils, water and sediment (Table 11-1).
No new information is available on the effects of Pb in saltwater algae at levels that are within the
concentrations of interest for this ISA (Section 11.1.1). There was, however, one new endpoint of note. In
a novel assay designed to assess the effects of toxicants on algal swimming behavior, Pb was shown to
inhibit motility in four saltwater algal motile species (Feng et al.. 2016). The lowest EC10 for 2-hour algal
swimming inhibition was 2.36 |iM (488 |ig Pb/L) in Platymoncts subcordiformis; effects on the three other
algal species tested were found at higher exposures. All exposures at which effects on swimming behavior
were observed support previous findings of Pb toxicity to algae at concentrations that greatly exceed
concentrations of Pb encountered in the natural environment.
There are a few studies on the effects of seepweeds (plants in the genus Suaeda found in
saltmarshes) which support previous findings of Pb toxicity at higher exposures. For instance, significant
negative effects on the growth of S. heteroptera were observed at concentrations of Pb higher than
400 mg Pb/kg (He et al.. 2016). A study of metabolic biomarkers in S. salsa revealed that Pb exposure at
20 |ig Pb/L could induce osmotic stress and disturbances in energy metabolism after long-term exposure
for 1 month, whereas no effects were seen in the short term (1 week) (Wu et al.. 2012b). Growth effects
were not seen in a congeneric species, S. fruitcosa, even at exposures of 600 (j,M (125 mg Pb/L), though
the study affirmed the plant nutrient content and activity of antioxidant enzymes were affected by metal
stress at high levels of exposure (Bankaii et al.. 2016). As in freshwater plants, Pb is concentrated in root
tissue, but sensitivity is species-specific. In general, effects in saltwater plants are observed at much
higher Pb exposures than are found in the natural environment.
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11.4.4.3 Effects on Saltwater Invertebrates
No studies reporting effects of Pb in saltwater invertebrates were reviewed in the 1977 Pb AQCD
or the 1986 Pb AQCD. In the 2006 AQCD, a few effects were noted in saltwater invertebrates including
gender differences in sensitivity to Pb in copepods, increasing toxicity of Pb with decreasing salinity in
mysids and effects on embryogenesis in bivalves (U.S. EPA, 2006). In the 2013 Pb ISA, available
evidence was sufficient to be suggestive of a causal relationship between Pb exposure and the endpoints
of physiological stress, hematological effects, reproduction, and development in saltwater invertebrates
(U.S. EPA, 2013). For all other effects, the evidence was inadequate to assess causality (Table 11-6). New
information for saltwater invertebrates since the 2013 Pb ISA includes additional studies that report
physiological perturbations associated with Pb exposure, including a few observations in previously
untested taxa. Only a few of the many studies identified in the literature search on suborganism-level
responses to Pb exposure in saltwater invertebrates were conducted in the low |ig Pb/L range and hence
met the criteria for inclusion in the ISA (Section 11.1.1).
11.4.4.3.1 Suborganism-Level Response
The majority of studies in saltwater invertebrates do not link the effects reported at the molecular
and cellular levels to effects at the organism level of biological organization (e.g., survival, growth,
reproduction). One study in Tiger prawn (Penaens monodon) exposed to a range of Pb concentrations (14
to 232 |ig Pb/L) in seawater for 30 days reported an increase in lipid peroxidation starting at the
56 |ig Pb/L exposure concentration. Chronic exposure yielded NOEC = 14 |ig Pb/L and
LOEC = 29 |ig Pb/L for survival in this species (Hariharan et al., 2012).
For the endpoint of physiological stress, many studies from the 2013 Pb ISA, especially those that
considered enzymatic responses to Pb exposure, were conducted with nominal Pb concentrations in
mollusks. Since the 2013 Pb ISA, additional studies have explored the mechanisms of Pb-induced
physiological stress in saltwater invertebrates by linking observed responses to changes in gene
expression. Over the course of a 4-week exposure of the mussel M. ednlis to 111.68 |ig Pb/L (0.54 |iM).
transcripts involved in the unfolding protein response were differentially expressed with Pb, which
correlated with the bioaccumulation of Pb in gill tissue (Poynton et al., 2014). In addition, a sequence of
unknown function showed a statistically significant relationship with Pb concentration in gill tissue, and
the authors proposed the sequence may be identified in the future as a dose-dependent Pb-specific
biomarker in this species.
Physiological stress responses associated with Pb exposure were also observed in sediment
bioassays with saltwater invertebrates. Biomarkers of oxidative stress, cellular damage and genotoxicity
were measured in the benthic bivalve A. trapezia following 56-day exposure to Pb-spiked sediments
(analytically verified concentration of 100 and 300 mg Pb/kg) (Taylor and Maher, 2012). Pb
concentration in bivalves (1 mg Pb/kg for the low concentration and 12 mg Pb/kg for the high
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concentration) suggested low bioavailability from sediment, especially at the lower concentration. The
total antioxidant capacity was statistically significantly reduced in both Pb treatments compared with the
control. Lysosomal stability in hepatopancreas of Pb-exposed bivalves was significantly decreased,
suggesting effects on cellular membrane integrity and function. In gill tissue, which is an important site
for Pb uptake in bivalves, there was a statistically significant increase in both treatment groups in
micronuclei frequency, a biomarker of genotoxicity. A similar suite of biomarkers was assessed in the
deposit-feeding bivalve T. deltoidalis, also exposed to Pb-spiked sediments (100 and 300 mg Pb/Kg)
(Taylor and Maher. 2014). In contrast to the filter-feeding A. trapezia, T. deltoidalis accumulated Pb to a
concentration equal to that of the spiked sediment over the course of the experiment (28 days). Exposed
T. deltoidalis individuals had significantly reduced total antioxidant capacity and significantly higher
lysosomal destabilization and micronuclei frequency compared with control organisms.
Evidence in the 2013 Pb ISA was suggestive of a causal relationship between Pb exposure and
hematological effects, primarily based on field studies that correlated ALAD activity to measured Pb
levels in bivalve tissue (Company et al.. 2011; Kalman et al.. 2008). Generally, these studies have noted
that Pb content varies significantly among species and is related to habitat and feeding behavior. A few
additional studies have reported hematological effects in Pb-exposed saltwater invertebrates; however, at
concentrations higher than the concentration of Pb typically encountered in seawater (Table 11-1). Duarte
et al. (2020) demonstrated various sublethal biomarkers were activated during 28-day exposure at a low
concentration of Pb (10.6 |ig Pb/L) in the crab U. cordatus. This concentration was too low to inhibit
ALAD activity in the crabs; however, metallothioneins were induced and DNA damage occurred in
exposed individuals. Various immunotoxic endpoints were assessed in hemolymph of the marine crab
Charybdis japonica during a 30-day exposure and subsequent quantification of Pb in tissue (Xu et al..
2019). At the lowest concentration, 0.066 (j,M (13.6 |ig Pb/L) immune responses were not significantly
different from control responses. At the next lowest concentration (0.132 (j,M, 27.2 |ig Pb/L), there was an
initial increase followed by a decrease in the total hemocyte count. Total hemocyte count was
significantly lower than control counts at the end of the exposure duration.
11.4.4.3.2 Organism-Level Response
Saltwater invertebrate studies that report effects on growth, reproduction and development, and
survival are primarily reviewed in the exposure-response section (Section 11.4.5). In sea hare (A.
califomica) exposed to Pb solely through diet over 2 or 3 weeks (green seaweed U. lactuca previously
exposed to analytically verified concentration of either 10 |ig Pb/L or 100 |ig Pb/L for 48 hours), growth
was significantly lower in the treatment groups compared with the control (Jarvis et al.. 2015).
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11.4.4.4 Effects on Saltwater Vertebrates
In the 2013 Pb ISA, there was inadequate evidence to infer causality relationships between Pb
exposure and effects in saltwater vertebrates (Table 11-6). Few studies on saltwater vertebrates were
reviewed in the 2013 Pb ISA or in the previous Pb AQCDs, especially for reproduction, growth, and
survival (endpoints that may have relevance to the population level of biological organization and higher).
Studies reviewed in the exposure-response section (Section 11.4.5, Table 11-7) of this appendix include
chronic toxicity data for growth and survival endpoints in saltwater fish species published since the 2013
Pb ISA. Summarized below are recent studies that report Pb perturbation on physiological endpoints in
fish and other saltwater vertebrates.
11.4.4.4.1 Fish
Most of the available studies in saltwater fish seek to identify molecular and cellular responses to
Pb exposure and do not report effects at the organism level of biological organization (e.g., survival,
growth, reproduction). Furthermore, studies since the 2013 Pb ISA that quantify effects on biomarkers in
saltwater and euryhaline fish are typically conducted at Pb concentrations considerably higher than
conditions found in natural environments. Nunes et al. (2014b) assessed the response of anadromous
European eel (A. angiiilla) to Pb exposure down to 165 |ig Pb/L in 28-day aqueous exposure studies and
observed no statistically significant effects on the biomarkers of neurotoxicity or peroxidative membrane
damage. Only gill tissue GST activity was significantly increased at 165 |ig Pb/L, and further increased
with higher Pb concentration. Similar 28-day chronic bioassays were performed in juvenile turbot
{Scophthalmns maximns). Very few significant effects were reported at the lowest concentration tested
(291 |ig Pb/L). Hepatic CAT activity significantly decreased, liver GST significantly increased and no
measurable changes in biomarkers of neurotoxicity were observed (Nunes et al.. 2014a). Fernandez et al.
(2015) evaluated the suitability of ALAD as a biomarker for Pb exposure in wild-caught red mullet
(Miillus barbatus) along several locations of the Spanish coast. Pb concentration in muscle tissue was
low. However, there was a weak, but significant, inverse relationship with ALAD activity; ALAD activity
showed no statistically significant relationship to the condition factor, gonadosomatic index and
hepatosomatic index of the fish.
Since the 2013 Pb ISA, a series of studies have further elucidated the effects of Pb exposure via
diet on multiple physiological endpoints in saltwater fish, and these perturbations were linked to a
decrease in weight gain (growth) in one study. In juvenile Korean rockfish (Sebastes schlegelii),
biomarkers of oxidative stress (SOD, GST) were significantly increased, AChE was significantly
decreased in muscle (Kim et al.. 2017). and physiological stress indicators (heat shock protein 70 mRNA
gene expression and plasma Cortisol) were significantly increased (Kim and Kang. 2016). as were
hematological parameters (hemocrit, hemoglobin) (Kim and Kang. 2015) by 4-week dietary
Pb > 60 mg/kg in experimental diet formulation. This is consistent with dietary exposure in starry
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flounder (Platichthys stellatus), in which the same hematological parameters as well as red blood cell
count were significantly decreased at 4-week dietary Pb exposure over 60 mg Pb/kg (Hwang et al.. 2016).
In rockfish, immune response was elicited at a higher dietary concentration (>120 mg Pb/kg) at 4 weeks
(Kim and Kang. 2016). A decrease in daily weight gain was observed in rockfish at >120 mg Pb/kg (Kim
and Kang. 2015). A dietary intake above 60 mg Pb/kg daily after 4 weeks of exposure to Pb appeared to
be the threshold for most effects.
11.4.4.4.2 Other Saltwater Vertebrates
There are a few new studies of nonfish saltwater vertebrates that report blood lead levels (BLLs)
and associated effects. In a survey of blood Pb levels in common eider ducks (Somateria mollissimct) at a
breeding colony in the northern Hudson Bay, birds with higher BLLs had lower body condition indexes
(body mass/head length) when they arrived at the breeding grounds (Provencher et al.. 2016). Birds with
higher BLLs arrived later at the breeding grounds. Birds that arrive later at the breeding grounds and with
lower body condition indexes are more likely to have lower reproductive success. A study of loggerhead
sea turtles (Cciretta caretta) in Casey Key, Florida examined the connection between blood Pb
concentrations and hematological effects (Perrault et al.. 2017). Over a range of blood Pb levels (0.07-
0.52 |ig/g dry weight), there was a significant negative relationship between BLLs and albumin, a2-
globulins, total solids, and Fe.
11.4.5 Exposure and Response of Saltwater Species
Evidence regarding exposure-response relationships and potential thresholds for Pb effects on
saltwater biota can provide tools for quantitative analyses of risks for coastal saltwater ecosystems. No
exposure-response studies in saltwater algae or vertebrates, and very few studies on saltwater
invertebrates, were reported in the 1977, 1986 or 2006 Pb AQCDs. For saltwater invertebrates, available
evidence at the time of the 2013 Pb ISA was suggestive of a causal relationship between Pb exposure and
reproductive and developmental effects (U.S. EPA. 2013). Much of the evidence was from exposure-
response bioassays.
Since the 2013 Pb ISA, new toxicity data for saltwater algae, invertebrates and fish have been
reported based on analytically verified Pb concentration. This information reduces uncertainties identified
in the previous review in terms of a lack of exposure-response data for saltwater biota, especially for
chronic toxicity, and enables calculations of effect levels for saltwater biota based on experimental data
(Church et al.. 2017). The studies listed in Table 11-7 are those that report exposure-response values at
concentrations comparable to, or lower than, the most sensitive saltwater biota identified in the 2013 Pb
ISA or the 2006 AQCD (i.e., the most environmentally relevant studies). Exposure-response data from
previously untested taxonomic groups are also discussed in this section. In general, marine organisms are
11-192
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tolerant of Pb at much higher concentrations than those encountered in uncontaminated natural
environments.
In studies reviewed in the 2013 Pb ISA, marine algae exhibited a range of sensitivity to Pb, with a
72-hour EC50 of 105 |ig Pb/L reported for Chaetorceros spp. Other tested species were considerably less
sensitive (72-hour EC50 = 740 |ig Pb/L or higher) (Debelius et al.. 2009). Exposure-response data for
marine algal species published since the 2013 Pb ISA greatly exceed environmental concentrations; for
example, in the marine alga Ncinnochloropsis oculata, the 72-hour IC50 = 1,810 (ig Pb/L for growth
inhibition (Zam an i - Ah m adm ah m oodi et al.. 2020). Longer-term exposure studies assessing the population
growth rates of polar marine algal species have reported effects as low as 24-day EC10 = 152 |ng Pb/L for
('ryolhecomonas armigera (Koppcl et al.. 2017) and a 10-day IC10 = 260 |ig Pb/L for Phaeocystis
antarctica (Gissi et al.. 2015).
In the 2013 Pb ISA, studies that reported effect concentrations in saltwater invertebrates included
a delay in reproduction onset in the marine amphipod, E. laevis, at 118 mg/Pb kg sediment, a
concentration the authors indicated was below the current marine sediment regulatory guideline for Pb
(218 mg Pb/kg sediment) (Ringenarv et al.. 2007; NOAA. 1999). A 96-hour EC50 = 197 |ig Pb/L for the
growth of larvae and EC50 = 297 |ig Pb/L for embryogenesis inhibition was observed for the clam
Meretrix meretrix (Wang et al.. 2009). Another study reported a decrease in the fertilization rate of eggs
of the marine polycheate Hydroides elegans-, in eggs pretreated with 48 |ig Pb/L, hatching decreased to
20% of control levels (Gopalakrishnan et al.. 2008). The lifestages of H. elegans varied in their sensitivity
to Pb, with the most sensitive period being larval settlement, with an EC50 of 100 |ig Pb/L. In the 2013 Pb
ISA, the most sensitive endpoint for growth in a saltwater invertebrate was LOAEL = 85 mg Pb/kg in
sediment in the polychaete Capitella sp. (Horng et al.. 2009). Other saltwater invertebrate exposure-
response studies in the 2013 Pb ISA reported effects at higher Pb concentrations. In the 2006 AQCD, the
most sensitive endpoint was a 48-hour EC50 = 221 |ig Pb/L and LOEC = 50 |ig Pb/L for embryogenesis in
the mussel M. galloprovincialis (based on nominal Pb concentration only) (Beiras and Albentosa. 2004).
Recent exposure-response data for saltwater invertebrates include reproductive and
developmental bioassay results based on analytically verified concentrations for mollusks and
echinoderms, with effects reported at lower concentrations than in studies included in the 2013 Pb ISA
(Table 11-7). Embryo development of the scallop Argopecten purpurcitus was impaired with Pb exposure,
with the 48-hour EC50 reported as = 44 |ig Pb/L (Romero-Murillo et al.. 2018). The order of sensitivity of
10 marine bivalve species (based on the percentage of normal D-veliger larvae assessed at 48 hours of Pb
exposure) was oysters > mussels > scallops > cockles > clams (Markich. 2021). The oystersM. gigas (48-
hour EC50 = 49.5 |ig Pb/L, 48-hour NEC = 9.9 |ig Pb/L) and S. glome rata (48-hour EC50 = 52.1 |ig Pb/L,
48-hour NEC = 10.1 |ig Pb/L) were most sensitive while the clam Iras crenatus (48-hour
EC50 = 196 |ig Pb/L, 48-hour NEC = 39.8 |ig Pb/L) was the least sensitive bivalve tested. In a series of
bioassays, Nadella et al. (2013) assessed Pb effects on embryo development in two mussels, M.
galloprovincialis and M. trossolus, and the sea urchin S. purpuratus. Both mussel species exhibited
11-193
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similar acute toxicity to Pb in 48-hour embryo-larval toxicity tests in 100% seawater (M.
gcilloprovincicilis-ECso = 63 |ig Pb/L, EC20 = 19 |ig Pb/L; EC10 = 10 (ig Pb/L, NOEC = 3.2 |ig Pb/L and
M. trossolus, ECso = 45 |ig Pb/L; EC20 = 16 |ig Pb/L; EC10 = 9 jig Pb/L; NOEC = 3.4 |ig Pb/L). In the 72-
hour embryo-larval toxicity test in the sea urchin S. purpurcitus, the EC50 = 74 |ig Pb/L,
EC20 = 31 |ng Pb/L, EC10 =19 |ng Pb/L and NOEC = 2.7 |ig Pb/L. In a similar 72-hour larval development
toxicity test with the sea urchin Evechimis chloroticus, the EC50 = 52.2 |ig Pb/L, with skeletal
abnormalities observed in the lower range of concentrations (10 (ig Pb/L and 20 |ig Pb/L) (Rouchon and
Phillips, 2017). However, Pb in the exposure water was not analytically verified in the study.
Developmental endpoints in oyster C. gigas were less sensitive to Pb, with EC50 = 660.3 |ig Pb/L for
embryo toxicity, 96-hour LC50 = 699.5 |ig Pb/L for larval mortality and LOEC = 96.7 |ig Pb/L for
significant increase of abnormal D-shaped larvae (Xie et al., 2017). In a series of fertilization bioassays
with the marine polychaete broadcast spawner Galeolarict caespitosa, the EC10 for reproduction varied
with the density of sperm used in the bioassays and ranged from 65 to 910 (ig Pb/L. The toxicity of Pb
was significantly decreased at higher sperm density (Lockyer et al., 2019). The EC10 was calculated to be
30 |ig Pb/L at a sperm density required to achieve 50% of the maximum fertilization.
New exposure-response data on previously untested marine invertebrate taxa, including species of
corals and sea anemones, generally show that these organisms are tolerant to Pb at relatively high
concentrations. Hedouin et al. (2016) assessed survival in adult and larval stages of the Scleractinian coral
Pocillopora damicornis. Results from 96-hour acute toxicity testing in adults collected during two
seasons near Oahu, Hawaii (summer 96-hour LC50 = 742 |ig Pb/L, winter 96-hour LC50 = 477 |ig Pb/L)
and coral larvae tested in the laboratory at two temperatures (96-hour LC50 = 681 |ig Pb/L at 27°C, 96-
hour LC50 = 462 |ig Pb/L at 30°C) showed similar tolerance to Pb. In Cnidarian (sea anemone) Aiptasia
pulchella, the 96-hour LC50 values were 8,060 |ig Pb/L and 12,400 |ig Pb/L in two separate tests. In the
same species, the 6-hour EC50 = 2,610 |ig Pb/L and 12-hour EC50 = 1,740 |ig Pb/L for rapid tentacle
retraction, suggesting that anemones are tolerant to Pb, even at concentrations that greatly exceed that of
Pb in seawater (Howe et al., 2014). In contrast, a 30-day exposure to Pb in the marine Tiger prawn P.
monodon yielded NOEC = 14 |ig Pb/L and LOEC = 29 |ig Pb/L for survival, suggesting that these
crustaceans are relatively sensitive to Pb (Hariharan et al., 2012). A recent review of Pb effects on marine
invertebrates by Botte et al. (2022) summarizes many of the effect concentrations and studies described
above.
For vertebrates, several studies published since the 2013 Pb ISA provide chronic toxicity data for
saltwater fish species, information that was previously lacking for evaluating the longer-term effects of Pb
on these organisms. Reynolds et al. (2018) conducted 28-day chronic toxicity tests with larval topsmelt
A. a ffinis (a fish species native to the coast of the western United States) at two salinities (14 ppt and 28
ppt) to represent conditions in estuarine and marine environments. In the larval fish, survival was affected
to a greater extent at the lower salinity (LC50 = 15.1 |ig Pb/L, NOEC <13.8 |ig Pb/L) than at the higher
salinity (LC50 = 79.8 |ig Pb/L, NOEC = 45.5 |ig Pb/L) due to the higher fraction of Pb in the form of Pb2+
at lower salinity. Growth effects (assessed as standard length) were reported in the same study, with
11-194
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greater response observed at the lower salinity (ECio = 16.4 |ig Pb/L) compared with the higher salinity
(ECio = 82.4 (ig/L). Tests conducted with juvenile topsmelt at 28 ppt (28-day LC50 = 167.6 |ig Pb/L)
showed that this lifestage was less sensitive to Pb than the larval stage (28-day LC50 = 79.8 |ig Pb/L). The
authors observed abnormal swimming and morphology, but these endpoints were not quantified.
Calculated chronic values for additional saltwater fish species include NOEC = 14 |ig Pb/L and
LOEC = 29 |ig Pb/L for grey mullet (M cephahis) fingerling survival and NOEC = 1 1 |ig Pb/L and
LOEC = 22 |ig Pb/L for Tiger perch (T. jarbna) fingerling survival following 30-day exposure to Pb
(Hariharan et al., 2016). The 96-hour LC50 values in these species were 2,570 and 2,990 |ig Pb/L,
respectively.
Given the increased availability of toxicity data for saltwater biota since development of the
AWQC for Pb by the U.S. EPA Office of Water in 1984 (U.S. EPA, 1985a) (Section 11.1.7.3), Church et
al. (2017) recently proposed updated U.S. saltwater acute AWQC of 100 |ig Pb/L (acute) and chronic
AWQC of 10 |ig Pb/L (chronic) based on genus mean toxicity values following U.S. EPA methodology
(U.S. EPA, 1985b). For the acute genus sensitivity distribution (Figure 11-6), data from 54 species and 49
genera were included. The proposed value of 100 |ig Pb/L is less than the current acute criterion of
210 (ig Pb/L due to toxicity data from relatively sensitive early lifestages of Echinodermata and Mollusca.
Although Church et al. (2017) derive regulatory values using SSD approaches, it is noted that some of the
toxicity values used in their analyses are from data sources not included in the IS As (i.e., unpublished
reports, university theses, memoranda).
11-195
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Diss. Pb, |jg/L
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Dissolved Pb and triangular distribution fit to the four lowest genus mean acute values following U.S. EPA guidelines. Genus mean
acute values (red circles); solid curved line = triangular distribution; dashed vertical line = final acute value; solid vertical
line = criterion maximum concentration (proposed acute criterion); black text = genera associated with genus mean acute values.
Source: Church et al. (2017).
Figure 11-6 Acute genus sensitivity distribution for saltwater biota from
Church et al. (2017).
The proposed Church et al. (2017) chronic value of 10 (ig Pb/L for saltwater (based on EC20 or, in
some cases, EC50 data divided by a factor of two when EC20 data could not be calculated from available
data) is based on data for 21 species and 17 genera. The four lowest genus mean chronic values were
10 |Lig Pb/L for a mysid, 28 |ig Pb/L for blue mussel (Mytilis spp.), 36 |ig Pb/L for purple sea urchin (S.
purpurcitus), and 55 |ig Pb/L for topsmelt (A. ctffinis). In their derivations of acute and chronic values,
Church et al. (2017) included some non-North American species. If the analysis was limited to North
American biota, the proposed acute and chronic values would be 110 (ig Pb/L and 8.8 (ig Pb/L,
respectively. Comparison of chronic sensitivity distributions in saltwater biota for dissolved Pb following
U.S. EPA and European Union methods is shown in Figure 11-7. Following the publication of these
proposed values, Reynolds et al. (2018) conducted additional testing with topsmelt larvae
(LC20 = 10.7 |ig Pb/L at 14 ppt salinity).
11-196
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Species mean chronic values (European Union method) are shown in red circles; genus mean chronic values (U.S. EPA method)
are shown in open circles; solid red curve = Weibull distribution fitted to species mean chronic values; solid black curve = triangular
distribution fit to the four most sensitive genus mean chronic values; dashed red vertical line = median 5th percentile hazardous
concentration based on Weibull distribution; dashed black vertical line = criterion continuous concentration (proposed chronic
criterion); black text = genera associated with genus mean chronic values; red text = species associated with species mean chronic
values.
Source: Church et al. (2017).
Figure 11-7 Comparison of chronic sensitivity distributions in saltwater biota
for dissolved Pb following the U.S. EPA and European Union
methods.
11-197
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Table 11-7 Studies in saltwater biota with analytically verified Pb concentration that report an effect on
growth, reproduction, or survival comparable to, or lower than, the lowest effect concentrations
reported in previous Pb AQCDs or the 2013 Pb ISA
Reference
Species Concentration Exposure Method ^Factors*' ^rKlooint Effect Concentration ''since the
P 2013 Pb
ISA)
Invertebrates
Mussel
(Mytilus
galloprovincialis)
Mussel
(Mytilus trossulus)
Sea urchin
(Strongylocentrotus
purpuratus)
Nominally 3.2, 10, 32,
100, 320,1,000 |jg Pb/L
(concentrations were
measured for each
individual assay)
Standard embryo development
acute toxicity tests for larvae of
mussel (to 48-hr
postfertilization) and sea urchin
(to 72-hr postfertilization)
conducted using ASTM
protocols in 100% sea water
DOC:
1.79 ± 0.02 mg/L
Additional toxicity
tests were
conducted with
added DOC
Salinity:
33 ppt
Developmental
assays
conducted over a
range of salinities
from 15 to 33 ppt
Temperature:
20°C ± 1°C
for mussels
15°C± 1°C
for sea urchin
Reproduction:
Development of
larvae: The
percentage of
embryos exhibiting
normal
development was
assessed after 48-
hr (mussels) or 72-
hr (sea urchin)
exposure to Pb at
varying
concentration in
seawater. The
acute toxicity of Pb
was similar
between the two
species of mussel
and sea urchin
M. galloprovincialis
48-hr ECso = 63 pg Pb/L
48-hr EC2o= 19 pg Pb/L
48-hr ECio= 10 pg Pb/L
48-hr
NOEC = 3.2 pg Pb/L
M. trossolus
48-hr ECso = 45 pg Pb/L
48-hr EC2o= 16 pg Pb/L
48-hr EC10 = 9 pg Pb/L
48-hr
NOEC = 3.4 pg Pb/L
S. purpuratus
72-hr ECso = 74 pg Pb/L
72-hr EC20 = 31 pg Pb/L
72-hr ECio= 19 pg Pb/L
72-hr
NOEC = 2.7 pg Pb/L
Nadella et
al. (2013)
Scallop
(Argopecten
purpuratus)
7 (control), 25, 50, 100,
140, 570, 730,1,000,
1,590 pg Pb/L
(measured)
48-hr embryo-larval
development assay with Pb-
nitrate conducted in 100% sea
water. In addition, a 96-hr
acute toxicity test was
Salinity:
35 ppt
pH:
Reproduction:
Embryos exhibited
abnormal
development
(impaired D-larvae
Embryo:
48-hr ECso = 44 pg Pb/L
Romero-
Murillo et
al. (2018)
Juvenile:
11-198
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Species
Concentration
Exposure Method
Modifying
Factors
Effects on
Endpoint
Effect Concentration
Reference
(Published
since the
2013 Pb
ISA)
conducted with juveniles
(21 mm in shell length)
8.0
Temperature
(embryo
exposure)
19°C ± 1°C
development)
following Pb
exposure.
Survival:
Assessed in
juvenile lifestage
only
96-hr
LCso = 1,420 pg Pb/L
Oyster
(Magallana gigas)
Oyster
(Saccostrea
glomerata)
Mussel
(.Xenostrobus
securis)
Scallop
(Scaeochlamys
livida)
Cockle
(Anadara trapezia)
Cockle
(Fulvia tenuicostata)
Each test with 1.5 to 2-hr-
old embryos (8-cell
stage) consisted of a
control and 12 metal
concentrations (based on
preliminary range-finding
tests). Concentrations
were analytically verified
but not reported
48-hr embryo-larval
development assay with Pb-
nitrate conducted in 100% sea
water. Test waters were not
renewed, and embryos were
not fed. Percentage of normal
D-veliger larvae was
determined by direct
observation of 100 larvae (per
replicate)
Salinity:
30 ppt ± 0.5%
pH:
7.85 ± 0.05
Temperature:
21°C±1°C
DO:
80% to 95%
saturation
Reproduction:
Embryos exhibited
abnormal
development
(impaired D-larvae
development)
following Pb
exposure. The
order of sensitivity
of the bivalves to
Pb was oysters >
mussels > scallops
> cockles > clams
M. gigas
48-hr
ECso = 49.5 |jg Pb/L
48-hr NEC = 9.9 pg Pb/L
S. glomerata
48-hr
ECso = 52.1 |jg Pb/L
48-hr
NEC = 10.1 |jg Pb/L
X. securis
48-hr
ECso = 59.9 |jg Pb/L
48-hr NEC = 12 pg Pb/L
S. livida
48-hr
ECso = 67.2 pg Pb/L
48-hr
NEC = 13.7 pg Pb/L
Markich
(2021)
Clam
(Hiatula alba)
Clam
A. trapezia
48-hr
ECso = 84.9 pg Pb/L
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Reference
Species Concentration Exposure Method ^Factors^ ^nd^oint Effect Concentration 'since the
^ on-10 du
2013 Pb
ISA)
(Barnea 48-hr
australasiae) NEC = 16.8 |jg Pb/L
Clam
(Spisula trigonella)
Clam
(Irus crenatus)
H. alba
48-hr
ECso = 129 |jg Pb/L
48-hr
NEC = 24.8 |jg Pb/L
B. australasiae
48-hr
ECso = 140 |jg Pb/L
48-hr NEC = 28 |jg Pb/L
S. trigonella
48-hr
ECso = 177 |jg Pb/L
48-hr
NEC = 36.7 |jg Pb/L
F. tenuicostata
48-hr
ECso = 108 |jg Pb/L
48-hr
NEC = 22.3 |jg Pb/L
I. crenatus
48-hr
ECso = 196 |jg Pb/L
48-hr
NEC = 39.8 |jg Pb/L
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Species
Concentration
Exposure Method
Modifying
Factors
Effects on
Endpoint
Effect Concentration
Reference
(Published
since the
2013 Pb
ISA)
Prawn
(Penaeus monodon)
1.7 (control-lab seawater
used in bioassays), 14,
29, 56, 108, 230 pg Pb/L
(measured)
Post larvae were exposed to
Pb acetate for 30 days in a
continuous flow-through
system. Prawns fed twice daily
and Pb concentrations
measured every 10 days
Salinity:
27.7 ± 0.5 ppt
Temperature:
25.4 ± 0.7°C
DO:
6.3 ± 0.6 mg/L
Survival:
Survival of P.
monodonwas
significantly
decreased at the
higher exposure
concentrations
30-d Hariharan
NOEC = 14 |jg Pb/L et al.
30-d LOEC = 29 pg Pb/L (2012)
pH: 7.1 ± 0.5
Vertebrates
28-d survival of larval
fish at 14 ppt salinity
LC5 = 7.7 pg Pb/L
LC10 = 8.3 pg Pb/L
LC15 = 9.9 pg Pb/L
LC20 = 10.7 pg Pb/L
LC25 = 11.5 pg Pb/L
LC40 = 13.6 pg Pb/L
LCso = 15.1 pg Pb/L
NOEC = <13.8 pg Pb/L
LOEC = 13.8 pg Pb/L
28-day survival of larval
fish at 28 ppt salinity
LC5 = 36.6 pg Pb/L
LC10 = 43.4 pg Pb/L
LC15 = 48.8 pg Pb/L
LC20 = 53.5 pg Pb/L
LC25 = 58.0 pg Pb/L
Topsmelt
(Atherinops affinis)
Measured: Mean ± SD
Low salinity, larval fish:
Total Pb
BDL, 17 ± 1 pg Pb/L,
34 ± 1 pg Pb/L,
69 ± 4 pg Pb/L,
85 ± 15 pg Pb/L,
127 ± 16 pg Pb/L
Dissolved Pb
BDL, 14 ± 1 pg Pb/L,
27 ± 2 pg Pb/L,
51 ± 3 pg Pb/L,
80 ± 7 pg Pb/L,
117 ± 19 pg Pb/L
High salinity, larval fish:
Total Pb
BDL, 58 ± 9 pg Pb/L,
107 ±20 pg Pb/L,
200 ± 14 pg Pb/L,
Larval fish (< 3 day old) were
tested in two different salinities
(14 ppt and 28 ppt) in 28-day
exposures to Pb nitrate
administered in a flow-through
test system set to replace the
total volume of synthetic
seawater in each 2-L exposure
chamber replicate once every
12 h. In addition, a 28-d
exposure was conducted with
juvenile fish (2.5 mo old) at
28ppt at higher Pb
concentration (control, 100 and
200 pg Pb/L)
Low Salinity
larval fish:
Salinity
14.1 ±0.1 ppt
Temperature:
18.2 ± 0.3°C
Alkalinity:
58 ± 5 mg/L as
CaC03
pH:
7.96 ± 0.17
DO:
7.58 ± 0.39
High Salinity
larval fish:
Salinity
28.1 ± 0.6 ppt
Temperature:
18.1 ± 0.2°C
Survival:
Pb was consistently
more toxic to larva
fish at the lower
salinity (14 ppt)
compared with the
higher salinity and
larvae were more
sensitive than
juvenile fish at
28 ppt. Free Pb2+
ion concentrations,
the most
bioavailable form of
Pb, were higher in
the lower salinity
water based on Pb
speciation
calculations.
Growth:
Growth effects in
larval fish
(assessed as
Reynolds
et al.
(2018)
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Species
Concentration
Exposure Method
Modifying
Factors
Effects on
Endpoint
Effect Concentration
Reference
(Published
since the
2013 Pb
ISA)
386 ± 43 |jg Pb/L,
563 ± 45 |jg Pb/L
Dissolved Pb
BDL, 46 ± 10 |jg Pb/L,
90 ± 20 |jg Pb/L,
171 ±22 |jg Pb/L,
259 ± 24 |jg Pb/L,
435 ± 48 |jg Pb/L
High salinity juvenile fish:
Total Pb
BDL, 154 ±67 pg Pb/L,
239 ± 98 |jg Pb/L
Dissolved Pb
BDL, 100 ±21 |jg Pb/L,
190 ± 30 |jg Pb/L
Alkalinity:
105 ± 8 mg/L as
CaC03
pH:
7.92 ± 0.07
DO:
6.88 ± 0.60
standard length)
were more
pronounced at the
lower salinity
(EC10 = 16.4 pg
Pb/L) compared
with the higher
salinity
(EC10 = 82.4 pg
Pb/L)
LC40 = 70.8 pg Pb/L
LCso = 79.8 pg Pb/L
NOEC = 45.5 pg Pb/L
LOEC = 89.9 pg Pb/L
28-day survival of
juvenile fish at 28 ppt
salinity
LC5 = 105.3 pg Pb/L
LC10 = 110.9 pg Pb/L
LC15 = 116.8 pg Pb/L
LC20 = 123 pg Pb/L
LC25 = 129.5 pg Pb/L
LC40 = 151.2 pg Pb/L
LCso = 167.6 pg Pb/L
28-d EC10 for larval
growth (standard length)
at 14ppt
salinity = 16.4 pg Pb/L
28-d EC for larval
growth (standard length)
at 28 ppt
salinity = 82.4 pg Pb/L
Grey mullet
(Mugil cephalus)
Tiger perch
(Terapon jarbua)
7, 16, 34, 65,
136 pg Pb/L (M.
cephalus)
7,15,29,60,118 pg Pb/L
(T. jarbua)
Wild-caught fingerlings (3.0-
4.5 cm in size) were acclimated
to laboratory conditions then
exposed to Pb as Pb acetate in
a continuous flow-through
system for 30d
Salinity:
33.5 ± 1.4 ppt
Temperature:
23.5 ± 0.9°C
pH:
7.8 ± 0.5
Survival:
Survival of M.
cephalus and T.
jarbua decreased
with the increase in
exposure
concentrations
Grey mullet:
30-d
NOEC = 14 pg Pb/L
30-d LOEC = 29 pg Pb/L
Tiger perch:
Hariharan
et al.
(2016)
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Species
Concentration
Exposure Method
Modifying
Factors
Effects on
Endpoint
Effect Concentration
Reference
(Published
since the
2013 Pb
ISA)
DO:
6.5 ± 0.6
30-d
NOEC = 11 |jg Pb/L
30-d LOEC = 22 pg Pb/L
54 species and 49
genera of
invertebrates and
fish (acute toxicity
data included in
derivation of
proposed updated
acute saltwater
quality criterion for
Pb)
21 species and 17
genera of
invertebrates and
fish (chronic toxicity
data included in
derivation of
proposed updated
chronic saltwater
quality criterion for
Pb)
Pb was analytically
verified in all studies
U.S. EPA guidelines (U.S.
EPA. 1985b) were used to
identify acceptable studies.
Acute: All included assays
were embryo-larval toxicity
studies reporting 48 to 96-hr
ECsos. The four lowest genus
mean acute values
(Strongylocentrotus
purpuratus = 75 pg Pb/L;
Mytilus spp = 123 pg Pb/L;
Paracentrotus
lividus = 363 pg Pb/L; and
Dendraster
excentricus = 371 pg Pb/L) and
a total of 49 genus mean
values were used to determine
a final acute value of
203.6 pg Pb/L. This value was
divided by 2 to derive the
proposed acute criterion based
on U.S. EPA methods.
Chronic: Based on EC20 from
lifecycle tests or EC50 data
divided by a factor of 2 when
EC20 data could not be
calculated and augmented with
48-hr toxicity data in some
cases. The four lowest genus
mean chronic values
(.Americamysis
bahia = 10 pg Pb/L; Mytilus
spp. = 28 pg Pb/L;
Acute toxicity
endpoints
included survival,
immobilization,
and embryo-larval
development
The proposed
updated acute
criterion is lower
than the current
U.S. EPA acute
saltwater criterion
of 210 pg Pb/L due
to embryo-larval
toxicity tests with
sensitive
echinoderm and
mussel species.
Chronic toxicity
endpoints
included survival,
growth,
development, and
reproduction
The proposed
updated chronic
criterion is greater
than the current
U.S. EPA acute
saltwater criterion
of 8.1 pg Pb/L.
Uncertainty in the
derivation of the
Proposed Saltwater
Acute Water Quality
Criterion: 100 pg Pb/L.
Limiting the derivation to
North American species,
the proposed criterion is
110 pg Pb/L.
Proposed Saltwater
Chronic Water Quality
Criterion: 10 pg Pb/L.
Limiting the derivation to
North American species,
the proposed criterion is
8.8 pg Pb/L
Church et
al. (2017)
11-203
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Reference
Species Concentration Exposure Method ^Factors^ ^nd^oint Effect Concentration 'since the
P 2013 Pb
ISA)
Strongylocentrotus
purpuratus = 36 |jg Pb/L;
Atherinops
affinis = 55 |jg Pb/L) and a total
of 17 genus mean values were
used to identify a chronic 5th
percentile of 10 |jg Pb/L
following U.S. EPA guidelines
chronic criterion
has decreased due
to increased
availability of
studies; an acute-
to-chronic ratio was
not used.
ASTM = American Society for Testing and Materials; BDL = below the method detection limit; CaC03 = calcium carbonate; DO = dissolved oxygen; DOC = dissolved organic carbon;
EC = effect concentration; LC = lethal concentration; LOEC = lowest observed effect concentration; mo = month(s); NEC = no-effect concentration; NOEC = no-observed-effect
concentration; Pb = lead.
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11.4.6
Saltwater Community and Ecosystem Effects
As discussed in the 1986 Pb AQCD (U.S. EPA, 1986), the 2006 Pb AQCD (U.S. EPA, 2006) and
the 2013 Pb ISA (U.S. EPA, 2013), the body of evidence was inadequate to infer a causal relationship
between Pb exposure and saltwater community- and ecosystem-level effects. Observations from field
studies in the 2006 Pb AQCD and the 2013 Pb ISA found either negative or null relationships between Pb
and species abundance, richness, and diversity in saltwater macroinvertebrates; however, Pb was not the
only contaminant in most observational studies, making it difficult to separate the effects of Pb from those
of other metal pollutants. New studies published since the 2013 Pb ISA examined the relationship
between Pb in sediment and saltwater as well as the community effects. Several reported negative or null
relationships between sediment Pb concentrations and foraminiferal abundance and community structure,
while others reported positive associations.
Foraminiferal diversity and community structure via changes in the abundance of certain taxa
have been found to vary with sediment Pb along environmental gradients in various locations including in
the Pearl River estuary, China (Li et al., 2013), the Ria de Aveiro lagoon, Portugal (Martins et al., 2011),
the San Jose Bay estuary, Puerto Rico (Martinez-Colon et al„ 2018), the Gulf of Milazzo, Sicily, Italy
(Cosentino et al., 2013), the Strait of Malacca, Malaysia (Minhat et al„ 2020) and Chilika lagoon in India
(Barik et al., 2022).
In the Pearl River estuary, surface sediment OC, grain size and benthic foraminifera communities
were assessed (Li et al„ 2013). Mean ± S.D. sediment Pb concentrations in the study area were
36.98 ± 11.18 mg Pb/kg (range: 13.5-62.9 mg Pb/kg). Trace metal concentrations in the sediment of Pb,
Cu, Co, Cr, Ni, V, and Zn were negatively correlated with the Shannon-Weaver index, Fisher a index,
species richness, and abundance of certain foraminiferal species. The CCA demonstrated that Pb
explained 7.5% of variation in the foraminiferal community.
Some foraminifera taxa were found to positively correlate with bioavailable Pb in the channels of
Ria de Aveiro, Portugal, but diversity was unaffected by bioavailable Pb (Martins et al., 2011). The
concentrations of Pb in the sediment in the resistant mineralogical phase, adsorbed by clay minerals, and
associated with OM ranged from about 20 mg Pb/kg to 180 mg Pb/kg. There was a positive correlation
between total bioavailable concentrations of Pb in the sediment (the fraction absorbed by clay and OM
and coprecipitated with carbonates) and the abundance of miliolids, and bioavailable Pb was not
significantly correlated with the abundance of Ammonia tepida, Bnlimina spp., Bolivina spp., Haynesinct
germanica, Elphidium spp., agglutinated spp., and Shannon diversity index. CCA indicated that miliolids
and agglutinated species were correlated with Pb and Al. Principal components analysis suggested that
higher bioavailable concentrations of Pb in addition to As, Cd, Cu, Ni, and Zn generally lead to less
diverse foraminifera communities and that the agglutinated foraminifera and miliolids were more tolerant
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to Pb than other taxa examined. The authors noted that agglutinated foraminifera and miliolids were
typically concentrated near the lagoon mouth where Pb concentrations were higher.
In another example, Pb was negatively correlated with the abundance of certain foraminiferal
taxa, but not to diversity metrics in the San Jose Bay estuary, Puerto Rico (Martinez-Colon et al.. 2018).
Sediment Pb concentration ranged from 2 to 38 mg Pb/kg in the lagoon. Pb was significantly negatively
correlated with the relative abundance of Amphisteginct gibbosa, Archaias angulatus, Asterigerina
ccirincita, Discorbis, Elphidium crispum, Heterostegina depressct, Miliolinella, Quinqueloculina
agglutinins, and Triloculina bicarinata and positively correlated with the relative abundance of
Triloculina and Quinqueloculina agglutinans. Pb sediment concentration was not significantly correlated
with any of the other foraminiferal abundances or diversity indices such as species diversity, Shannon's
index, Equitability Index, foraminiferal density, or the relative abundances of Ammonia.
Pb enrichment factors were slightly positively correlated w ith Ammonia spp. {Ammonia beccarii,
A. gaimardii, A. tepida, and A. parkinsoniana) and low-oxygen foraminiferal assemblages in the Gulf of
Milazzo, Italy, but not to total deformed foraminifera, foraminiferal density, or the abundance of other
foraminiferal taxa (Coscntino et al.. 2013). Pb concentrations in the sediment ranged from 4.75 to
49.19 mg Pb/kg. Finally, Pb and Al were negatively correlated with foraminiferal abundance across a
gradient of sites in the Strait of Malacca, Malaysia (Minhat et al.. 2020). with Pb showing the greatest
enrichment among all metals, with values ranging from 8.8 to 29.2 mg Pb/kg. Overall, dissolved oxygen,
depth, Al, and Pb concentrations explained the most variation in foraminiferal species distributions. The
abundance of Ammonia tepida, which was the highest, was not significantly correlated with Pb sediment
concentration, while those of Bulimina marginata, Pararotalia ozawai, and Nonion subturgidum were
negatively correlated with Pb.
Foraminiferal abundance and diversity were correlated with certain bioavailable Pb sediment
concentrations in Chilika, which is the largest brackish water lagoon in Asia (Barik et al.. 2022). The
concentrations of Pb in the sediment were 68.27 ± 22.14 mg Pb/kg (mean ± S.D.) across 22 sampling sites
(range: 22.14-107.57 mg Pb/kg). Pb was statistically significantly positively correlated to the
concentrations of Co. In addition to Pb concentrations in the sediment, bioavailable fractions of Pb and
other heavy metals were determined. Specifically, Pb in the first fraction is the Pb bound to carbonates,
the second fraction includes Pb bound to FeMn oxides, the third is bound to OM, and the fourth is bound
to silicate. Pb concentration was significantly negatively correlated to the percentage of Pb in the second
(reducible) and third (oxidizable) fractions and positively correlated to the percentage of Pb in the fourth
(residual) fraction. Pb concentration alone was not correlated to the total number of live and dead
abundance, diversity, or species richness, while the percentage of Pb in the first fraction was positively
correlated to the abundance of dead foraminifera per gram sediment and negatively correlated to the
diversity of dead foraminifera. The diversity, measured by the Shannon diversity index, of live and dead
foraminifera was negatively correlated to the percentages of Pb in the second and third fractions and
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positively correlated to the percentage of Pb in the fourth fraction. Finally, live and dead foraminiferal
species richness was significantly negatively correlated to the percentage of Pb in the third fraction.
In summary, some mesocosm and observational studies published since the 2013 Pb ISA found
reductions in foraminiferal and meiofaunal community richness, diversity or abundance associated with
higher concentrations of Pb in sediment and water (Barik et al.. 2022; Minhat et al.. 2020; Martinez-
Colon et al.. 2018). Other studies found positive or null correlations (Barik et al.. 2022; Martinez-Colon et
al.. 2018; Cosentino et al.. 2013; Martins et al.. 2011).
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United States
Environmental Protection
Agency
EPA/600/R-23/375
January 2024
www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 12: Process for Developing the Pb
Integrated Science Assessment
January 2024
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Enviromnental Protection Agency
-------
CONTENTS
DOCUMENT GUIDE 12-iii
LIST OF TABLES 12-v
LIST OF FIGURES 12-vi
ACRONYMS AND ABBREVIATIONS 12-vii
APPENDIX 12 PROCESS FOR DEVELOPING THE Pb INTEGRATED SCIENCE
ASSESSMENT 12-1
12.1 Introduction 12-2
12.2 Documentation 12-2
12.2.1. Literature Database: Health and Environmental Research Online 12-2
12.2.2. Study Quality Documentation: Health Assessment Workspace Collaborative 12-3
12.3 Overview of the Process Steps for Developing Integrated Science Assessments 12-3
12.4 Relevance and Scope 12-5
12.4.1. Atmospheric Sciences 12-5
12.4.2. Exposure, Toxicokinetics, and Biomarkers 12-6
12.4.3. Health 12-7
12.4.4. Welfare—Effects on Terrestrial and Aquatic Ecosystems 12-11
12.5 Literature Search 12-15
12.5.1. Title and Abstract Screening 12-17
12.6 Study Selection: Full-Text Screening and Evaluation of Studies 12-20
12.6.1. Individual Study Quality 12-20
12.7 Peer Review and Public Participation 12-28
12.7.1. Request for Information 12-28
12.7.2. Integrated Review Plan 12-29
12.7.3. Peer Input 12-29
12.7.4. Internal Technical Review and Clearance 12-30
12.7.5. Clean Air Scientific Advisory Committee Peer Review 12-30
12.8 Quality Assurance and Quality Control 12-32
12.9 Conclusion 12-32
12.10 References 12-33
12-iv
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LIST OF TABLES
Table 12-1 Population, Intervention, Comparison, Outcome, and Context statement to define the
parameters and provide a framework for identifying relevant atmospheric science studies_
12-6
Table 12-2 Population, Exposure, Comparison, Outcome, and Study Design statement to define the
parameters and provide a framework for identifying relevant experimental studies
12-8
Table 12-3 Population, Exposure, Comparison, Outcome, and Study Design statement to define the
parameters and provide a framework for identifying relevant epidemiologic studies 12-10
Table 12-4 Level of Biological Organization, Exposure, Comparison, Endpoint, and Study Design
statement to define the parameters and provide a framework for identifying relevant
ecological studies 12-13
Table 12-5 Scientific considerations for evaluating the strength of inference from studies on the health
effects of Pb 12-24
12-v
-------
LIST OF FIGURES
Figure 12-1 General process for developing Integrated Science Assessments. 12-4
Figure 12-2 Literature flow diagram forthe Pb Integrated Science Assessment. 12-17
12-vi
-------
ACRONYMS AND ABBREVIATIONS
AQCD
Air Quality Criteria Document
NHANES
BLL
blood lead level
CASAC
Clean Air Scientific Advisory
ORD
Committee
Pb
CI
confidence interval
PbB
FRN
Federal Register Notice
PECOS
GFR
glomerular filtration rate
HAWC
Health Assessment Workspace
Collaborative
PICOC
HERO
Health and Environmental Research
PM
Online
PQAPP
IQ
intelligence quotient
IRP
Integrated Review Plan
QA
ISA
Integrated Science Assessment
QAPP
LECES
Level of Biological Organization,
QC
Exposure, Comparison, Endpoint, and
RBC
Study Design
SWIFT-AS
LOD
limit of detection
U.S. EPA
NAAQS
National Ambient Air Quality
Standards
NASGLP
North American Soil Geochemical
Landscapes Project
National Health and Nutrition
Examination Survey
Office of Research and Development
lead
blood lead concentration
Population, Exposure, Comparison,
Outcome, and Study Design
Population, Intervention, Comparison,
Outcome, and Context
particulate matter
Program Quality Assurance Project
Plan
quality assurance
Quality Assurance Project Plan
quality control
red blood cell
SWIFT-Active Screener
United States Environmental Protection
Agency
12-vii
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APPENDIX 12 PROCESS FOR DEVELOPING THE Pb
INTEGRATED SCIENCE ASSESSMENT
Summary of Public Resources for the 2024 Pb ISA
This appendix describes the process for developing the Lead (Pb) Integrated Science
Assessment (ISA), including literature search and screening methods; peer input and peer review; and
public participation. This table summarizes the publicly available resources related to this ISA and its
development. Readers looking for Federal Register Notices (FRNs) may search http://
www.regulations.gov by either the document citation number (the reference number to the specific
FRN) or the Docket ID number (reference number for the overall docket that may house multiple
FRNs, as well as public comments in response to those FRNs).
2024 Pb ISA
https://assessments.epa. aov/isa/document/&deid=359
536
Clean Air Scientific Advisory
Committee
https://casac.epa.aov/ords/sab/f?p=113:1
Federal Register Notices
http://www.requlations.qov
Document Citation: 85 FR 40641
Request for Information
Docket ID: EPA-HQ-OAR-2020-0312-0001
Integrated Review Plan,
Document Citation: 87 FR 13732
Volume 2
Docket ID: EPA-HQ-OAR-2020-0312-0010
Document Citation: 87 FR 27147
Peer Input Workshop
Docket ID: EPA-HQ-ORD-2020-0701-0001
Pb ISA External Review
Document Citation: 88 FR 19302
Draft
Docket ID: EPA-HQ-ORD-2020-0701
Integrated Review Plan
https://www.epa.qov/naaqs/lead-pb-standards-
planninq-documents-current-review
ISA Preamble
https://cfpub.epa.qov/ncea/isa/recordisplav.cfm?deid=
310244
Literature
https://hero.epa.qov/hero/index.cfm/proiect/paqe/proie
ct id/4081
Peer Input Workshop
https://cfpub.epa.qov/ncea/isa/recordisplav.cfm?deid=
354420
Study Quality Evaluations
https://hawc.epa.qov/assessment/100500318/
12-1
-------
12.1 Introduction
Integrated Science Assessments (ISAs) provide the scientific foundation for the review of the
primary (health-based) and secondary (welfare1-based) National Ambient Air Quality Standards
(NAAQS). ISAs contain a synthesis and evaluation of the most policy-relevant science using methods and
approaches described in the Preamble to the Integrated Science Assessments (U.S. EPA. 2015b). hereafter
"Preamble," which provides an overview of the ISA development process. The 2024 Pb ISA builds upon
the conclusions and scientific evidence from the 2013 Pb ISA (U.S. EPA. 2013a) and prior Air Quality
Criteria Documents (AQCDs) for Pb from 1977 (U.S. EPA. 1977). 1986 (U.S. EPA. 1986). and 2006
(U.S. EPA. 2006). and includes recent literature published since September 2011, the literature cutoff date
of the 2013 Pb ISA. In March 2022, the United States Environmental Protection Agency (U.S. EPA)
released the first two volumes of the Integrated Review Plan (IRP) for the Pb NAAQS review. Volume 2
of the IRP (U.S. EPA. 2022) identifies policy-relevant issues (i.e., those intended to frame the review and
focus it on the critical scientific and policy questions related to the adequacy of the standards) and
describes key considerations in the U.S. EPA's development of the Pb ISA. Volume 2 was made available
for public comment and a consultation with the U.S. EPA's Clean Air Scientific Advisory Committee
(CASAC) Pb Review Panel at a public meeting on April 8. 2022. The 2024 Pb ISA has been developed
by U.S. EPA scientists in the Office of Research and Development (ORD), other U.S. EPA scientists with
relevant experience, and external authors from ICF, a U.S. EPA contractor. The general ISA development
steps are presented in Figure 12-1, though particular details can vary across assessments. This appendix
supplements the 2015 ISA Preamble (U.S. EPA. 2015b) and Volume 2 of the IRP (U.S. EPA. 2022). and
further describes the process of developing the 2024 Pb ISA, including methods for documentation,
literature review, study quality evaluation, public engagement, and quality assurance (QA).
12.2 Documentation
12.2.1. Literature Database: Health and Environmental Research Online
To improve transparency, studies considered in the development of the ISAs are documented in
the U.S. EPA Health and Environmental Research Online (HERO) database. The publicly accessible
HERO project page for the 2024 Pb ISA contains the references that were considered for inclusion and
provides bibliographic information and abstracts. Within HERO, each reference has a unique HERO ID
'Under The Clean Air Act section 302(h) (42 U.S.C. § 7602(h)), effects on welfare include "effects on soils, water,
crops, vegetation, manmade materials, animals, wildlife, weather, visibility, and climate, damage to and
deterioration of property, and hazards to transportation, as well as effects on economic values and on personal
comfort and well-being."
12-2
-------
number. References can be viewed individually or filtered by appendix, discipline, or the draft in which
they are referenced.
Inclusion and exclusion decisions for references at each stage of screening are recorded by a
tagging system and are documented in the HERO database. A two-step screening process (title and
abstract screening and full-text screening) was used for this ISA; subsequent sections of this appendix
discuss the screening process in greater detail. References that passed through title and abstract screening
are tagged in HERO as "Title-Abstract Screening Included." Inclusion and exclusion decisions from full-
text screening of references passing through title and abstract screening are tagged in HERO as "Full-Text
Screening Included." References identified from sources other than literature searches were also screened
using the same discipline-specific criteria, and inclusion and exclusion decisions for these references are
also documented in HERO. Specific data about concentrations, experimental design, and results are
reported within the appendices.
12.2.2. Study Quality Documentation: Health Assessment Workspace
Collaborative
Reference-specific information about study quality is documented in the U.S. EPA Health
Assessment Workspace Collaborative (HAWC) for select health studies and can be accessed through the
HAWC project page for this ISA. All decisions about full-text screening are additionally documented in
the HERO database and on the publicly available HERO project page for this ISA. See Section 12.6 for a
more detailed discussion about study quality.
12.3 Overview of the Process Steps for Developing Integrated
Science Assessments
As described in the Preamble and shown in Figure 12-1, developing an ISA consists of the
following steps: literature search and study selection; evaluating study quality; developing initial draft
materials for peer-input consultation; evaluating, synthesizing, and integrating evidence; and developing
scientific conclusions and causality determinations (U.S. EPA. 2015b).
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Literature Search and
Study Selection
*
Evaluation of Individual Study Quality
After study selection, the quality of individual studies is evaluated by EPA or outside experts in the fields of
atmospheric science, exposure assessment, dosimetry, animal toxicology, controlled human exposure studies,
epidemiology, ecology, and otherwelfare effects, considering the design, methods, conduct, and documentation of
each study. Strengths and limitations of individual studies that may affect the interpretation of the study are
considered.
*
Develop Initial Sections
Review and summarize new study results as well
as findings and conclusions from previous
assessments by category of outcome/effectand
by discipline, e.g., toxicological studiesoflung
function.
Peer Input Consultation
Review of initial draft materials by scientists
from both outside and within EPA in public
meeting orpublic teleconference.
*
Evaluation, Synthesis, and Integration of Evidence
Integrate evidence from scientific disciplines - for example, toxicological, controlled human exposure, and
epidemiologic study findings for a particular health outcome. Evaluate evidence for related groups of endpoints or
outcomes to draw conclusions regarding health orwelfare effect categories, integrating health orwelfare effects
evidence with information on mode of action and exposure assessment.
Development of Scientific Conclusions and Causal Determinations
Characterize weight of evidence and develop judgments regarding causality for health or welfare effect categories.
Develop conclusions regarding concentration- or dose-response relationships, potentially at-risk populations,
lifestages, or ecosystems.
it
Draft Integrated Science Assessment
Evaluation and integration of newly published studies
after each draft.
Clean Air Scientific Advisory Committee
Independent review of draft documents for scientific
quality and sound implementation of causal
framework; anticipated review of two drafts of ISA in
public meetings.
Public Comments
Comments on draft ISA solicited by EPA
Final Integrated Science Assessment
Source: Modified from Figure II of the Preamble to the Integrated Science Assessment (U.S. EPA. 2015b).
Figure 12-1 General process for developing Integrated Science Assessments.
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12.4 Relevance and Scope
As a synthesis and evaluation of the most policy-relevant science, the 2024 Pb ISA includes
information on atmospheric science, exposure assessment, experimental health studies, epidemiologic
health studies, and studies of effects on terrestrial and aquatic ecosystems. For the 2024 Pb ISA, "policy-
relevant" science is described in Volume 2 of the IRP (U.S. EPA. 2022) as referring to "scientific
information and analyses intended to address key questions related to the adequacy of the standards."
Those "key questions" are also laid out in Volume 2 of the IRP. As stated in the Preamble (U.S. EPA.
2015b). "The key policy-relevant questions included in the IRP serve to clarify and focus the NAAQS
review on the critical scientific and policy issues, including addressing uncertainties discussed during the
previous review and newly emerging literature." The sections below describe the approaches and scoping
statements used to identify relevant studies in each discipline. The use of scoping statements to define
study relevance is consistent with recommendations by the National Academies of Sciences, Engineering,
and Medicine for improving the design of risk assessment through planning, scoping, and problem
formulation to better meet the needs of decision makers (NASEM. 2018).
12.4.1. Atmospheric Sciences
Studies were considered relevant for inclusion in the 2024 Pb ISA if they were judged to provide
original data and to substantially advance the understanding of Pb emission sources; atmospheric and
environmental processes (including chemistry and transport); measurement and estimation methods; or
recent concentrations and trends. This approach to determining study relevance required judgments about
whether a subject area of the research had the potential to inform policy specific to the NAAQS, and
whether a study published in the area provided sufficiently original results to add to the existing body of
knowledge.
Table 12-1 shows the relevance criteria used for broadly identifying recent environmental
research advances and knowledge gaps. These criteria are based on the approach described by Mengist et
al. (2020). who formulated a Population, Intervention, Comparison, Outcome, and Context (PICOC)
statement that designated the population as the population of scientific research work itself and the
outcome as the assessment of its knowledge and gaps.
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Table 12-1 Population, Intervention, Comparison, Outcome, and Context
statement to define the parameters and provide a framework for
identifying relevant atmospheric science studies
Concept Application
Population Include policy-relevant scientific research on Pb source emissions, environmental processes
(including chemistry and transport), measurement and estimation methods, and concentration
and trends.
Intervention Assess policy-relevant scientific advances and knowledge gaps.
Comparison Evaluate emissions, concentrations, and their rates of change across sources, atmospheric and
environmental processes, measurement and estimation methods, long-term temporal scales,
seasons, diurnal cycles, geographic regions, and urban and neighborhood spatial scales.
Outcome Identify policy-relevant scientific advances and knowledge gaps.
Context Focus on policy-relevant research performed in the United States or Canada; for some topics,
research performed outside of the United States or Canada can be excluded if sources or
concentrations are not relevant to the United States or if the body of research is very large; for
other topics, if source and concentration differences are not relevant to the topic or the number of
publications is very small, non-U.S. research can be included.
Pb = lead.
12.4.2. Exposure, Toxicokinetics, and Biomarkers
The following guidelines were used to judge the relevance of studies examining Pb exposures,
toxicokinetics, and biomarkers. Studies were included if they provided original data and substantially
advanced understanding of Pb exposure through environmental media and other pathways; Pb
toxicokinetics including uptake, distribution, metabolism, and elimination from the body; Pb biomarker
measurement techniques; Pb biomarker concentration trends; and the relationships between Pb in
environmental media and Pb biomarker concentrations, including biokinetic and empirical modeling of
those relationships.
Exposure studies pertaining to the U.S. population and U.S.-based Pb sources were preferred.
Studies were included from outside the United States if these studies were judged to have important
findings, with a focus on studies from Canada, western Europe, and Australia (i.e., areas with study
populations and air quality characteristics most similar to the United States). If it was deemed that studies
from the United States, Canada, western Europe, or Australia were not adequate (i.e., little to no
information that advanced understanding of a particular topic was found), then it was necessary to
consider all studies regardless of geographic location. For Pb toxicokinetics and biomarker measurement
techniques, studies, regardless of geographic location, were considered since the physical location in
which a study took place may have less bearing on results. Finally, although exposures in relation to Pb in
ambient air and originating from air-related sources are the focus of the appendix, studies containing Pb
concentrations in other media (soil, dietary sources, consumer products, occupational sources, and
ammunition) were included because cumulative body burden can occur as a result of contributions from
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multiple exposure pathways (e.g., ingestion of Pb-containing soil by children) and the origin of Pb can be
difficult to determine as stemming from an air-related source.
12.4.3. Health
Relevance for studies that evaluate the relationship between Pb exposure and health effects was
assessed using scoping statements that define the relevant Population, Exposure, Comparison, Outcome,
and Study Design (PECOS). Discipline-specific PECOS statements for epidemiologic and experimental
studies (i.e., animal toxicology studies) were developed to establish inclusion criteria based on the
objectives of the review, facilitating identification of the most relevant literature to inform the Pb ISA
(Table 12-2 and Table 12-3). In some cases, PECOS statements differ by health outcome depending on
well-established areas of research; gaps in the literature; and inherent uncertainties in specific
populations, exposure metrics, comparison groups, and study designs identified in the 2013 Pb ISA.
Additionally, some epidemiologic PECOS statements were further refined to emphasize the strongest
recent epidemiologic studies that address key uncertainties from the previous review; these PECOS
refinements are identified and described in detail in the relevant appendices. The use of PECOS
statements is widely accepted and often applied in the health disciplines for systematic review in risk
assessment. PECOS statements for the 2024 Pb ISA can also be found in each health effects appendix.
12.4.3.1. Experimental Studies
For experimental studies (specifically animal exposure studies), the relevance evaluation focused
on studies with appropriate study designs and relevant exposure concentrations (Table 12-2). The scope
of the experimental evidence used for the 2024 Pb ISA encompassed studies of nonhuman mammalian
animal species with exposures that are relevant to the range of human exposures (blood Pb levels [BLLs]
up to 30 (ig/dL, which is about one order of magnitude above the 95th percentile of the 2011-2016
National Health and Nutrition Examination Survey [NHANES] distribution of BLLs in children) (Eganet
al.. 2021).
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Table 12-2
Population, Exposure, Comparison, Outcome, and Study Design
statement to define the parameters and provide a framework for
identifying relevant experimental studies
Concept
Application
Population
Laboratory nonhuman mammalian animal species (i.e., mouse, rat, Guinea pig, minipig, rabbit,
cat, dog; whole organism) at any lifestage (including preconception, in utero, lactation,
peripubertal, and adult stages).
Exposure
Oral, inhalation, or intravenous routes administered to a whole animal (in vivo) that results in a
BLL of 30 pg/dLor below.ab
Comparison
A concurrent control group exposed to vehicle-only treatment or untreated control.
Outcome
Cancer and noncancer health outcomes including cardiovascular, dermal, developmental,
endocrine system, gastrointestinal, hematological, hepatic, immunological, metabolic syndrome,
musculoskeletal, neurological, ocular, renal, reproductive, or respiratory effects.
Study
Controlled exposure studies of animals in vivo.
Design
BLL = blood lead level; Pb = lead.
aPb mixture studies are included if they employ an experimental arm that involves exposure to Pb alone.
This level is approximately an order of magnitude above the upper end of the distribution of U.S. young children's BLLs. The
95th percentile of the 2011-2016 NHANES distribution of BLL in children (1-5 years; n = 2,321) is 2.66 |jg/dL (Egan et al.. 2021).
and the proportion of individuals with BLLs that exceed this concentration varies depending on factors including housing age,
geographic region, and a child's age, sex, and nutritional status.
12.4.3.2. Epidemiologic Studies
To identify the most relevant epidemiologic literature, the body of evidence from the 2013 Pb
ISA was considered in the development of the PECOS statements. Specifically, the scope of the current
assessment is informed by well-established areas of research, gaps in the literature, inherent uncertainties
in specific populations, exposure metrics, comparison groups, and study designs identified in the 2013 Pb
ISA. The evaluation of epidemiologic studies focused on the association between exposure to Pb (as
indicated by Pb levels in blood, bone, and teeth; validated environmental indicators of Pb exposure; or
intervention groups in randomized trials and quasi-experimental studies) and an ensemble of health
effects, including effects on the nervous system, cardiovascular effects, and reproductive and
developmental outcomes (Table 12-3). Emphasis was placed on studies conducted in non-occupationally
exposed populations, but recent longitudinal studies of occupational exposure to Pb published since the
literature cutoff date for the 2013 Pb ISA were considered insofar as they addressed atopic that was of
particular relevance to the NAAQS review (e.g., longitudinal studies designed to examine recent versus
historical Pb exposure). Additionally, the following types of epidemiologic studies are generally
considered to fall outside the scope and are not included in the ISA: review articles (which typically
present summaries or interpretations of existing studies rather than bringing forward new information in
the form of original research or new analyses); Pb poisoning studies or clinical reports (e.g., involving
accidental exposures to very high amounts of Pb described in clinical reports that may be extremely
unlikely to be experienced under ambient air exposure conditions); and risk or benefit analyses (e.g., that
apply existing concentration-response functions or effect estimates to exposure estimates for differing
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cases). Although review articles are not typically included in the health sections of the ISA, they are
identified and tracked during the literature searching and study selection phase of the assessment. These
reviews are often consulted to ensure that all relevant literature has been identified and to track key issues
related to a particular evidence base.
For some health outcomes for which the evidence assessed in the 2013 Pb ISA supported a
"causal" relationship, the epidemiologic PECOS statements were refined in order to further emphasize the
strongest recent epidemiologic studies that address the key uncertainties from the previous review and the
scientific questions in Volume 2 of the IRP (U.S. EPA. 2022). These PECOS refinements, which are
identified and described in detail in the relevant appendices, generally focus on the most informative
study designs and relevant BLLs, and emphasize control for important potential confounders that were
identified in the 2013 Pb ISA. Studies that met the broader PECOS criteria in Table 12-3, but were no
longer relevant under the refined criteria were still included in evidence inventories that summarize key
study details, including study population, exposure assessment, confounders, and select results.
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Table 12-3 Population, Exposure, Comparison, Outcome, and Study Design statement to define the parameters
and provide a framework for identifying relevant epidemiologic studies
Population: Any human population, including specific populations or lifestages that might be at increased risk of a health effect.
Exposure: Exposure to Pba as indicated by biological measurements of Pb in the body, with a specific focus on Pb in blood, bone, and teeth; validated environmental
indicators of Pb exposure, or intervention groups in randomized trials and quasi-experimental studies.
Comparison: Populations, population subgroups, or individuals with relatively higher versus lower levels of the exposure metric (e.g., per unit or log unit increase in the
exposure metric, or categorical comparisons between different exposure metric quantiles).
Outcome
Nervous System
Cardiovascular
Renal
Immune
Hematological
Reproductive
Developmental
Cancer
Other
Nervous system
Cardiovascular
Renal
Immune system
Hematological
Reproductive
Developmental
Cancer
Effects on the
effects including
effects including
effects
effects including
effects including
effects, including
effects, including
incidence,
hepatic system,
cognitive function
coronary heart
including
immunotoxicity,
disruption of
altered age of
adverse
mortality, or
gastrointestinal
(e.g., IQ decrement),
disease,
elevated
systemic
heme synthesis
puberty onset,
pregnancy
related
system,
externalizing and
hypertension
serum
inflammation, and
and RBC
reduced fertility,
outcomes (e.g.,
biomarkers.
endocrine
internalizing
and increased
creatinine
immune-based
function.
poor semen
reduced fetal
system, bone
behaviors,
blood pressure,
levels and
diseases.
quality or
growth, preterm
and teeth, ocular
psychopathological
and
lower
motility, and
birth, small for
health, and
effects, sensory
cardiovascular-
GFR.
miscarriage.
gestational age,
respiratory
organ function, motor
related mortality.
birth defects), as
system.
function, and
well as postnatal
neurodegenerative
developmental
diseases.
effects.
Study Design: Epidemiologic studies consisting of longitudinal and retrospective cohort studies, case-control studies, cross-sectional studies with appropriate timing of
exposure for the health endpoint of interest, randomized trials, and quasi-experimental studies examining interventions to reduce exposures.
GFR = glomerular filtration rate; IQ = intelligence quotient; Pb = lead; RBC = red blood cell.
aThe focus was on populations with nonoccupational Pb exposures, though recent longitudinal studies of occupational exposure to Pb were considered insofar as they addressed a topic that
was of particular relevance to the NAAQS review (e.g., longitudinal studies designed to examine recent versus historical Pb exposure).
bStudies that estimate Pb exposure by measuring Pb concentrations in particulate matter with a nominal mean aerodynamic diameter less than or equal to 10 |jm (PM10) and particulate matter
with a nominal mean aerodynamic diameter less than or equal to 2.5 |jm (PM25) ambient air samples are only considered for inclusion if they also include a relevant biomarker of exposure (e.g.,
Pb in blood, bone, or teeth). Given that size distribution data for Pb-PM are fairly limited, it is difficult to assess the representativeness of these concentrations to population exposure [Section
2.5.3 (U.S. EPA. 2013a1l. Moreover, data illustrating the relationships of Pb-PM10 and Pb-PM2.5 with blood Pb levels are lacking.
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12.4.4. Welfare—Effects on Terrestrial and Aquatic Ecosystems
For welfare effects (i.e., on terrestrial and aquatic ecosystems), scoping statements defining the
Level of Biological Organization, Exposure, Comparison, Endpoint, and Study Design (LECES) were
used. U.S. EPA developed the LECES based on the PECOS with some concepts substituted to provide a
better fit with ecological science. In the LECES, "population" (PECOS) is replaced with "level of
biological organization" (LECES) and "outcome" (PECOS) is replaced with "endpoint" (LECES). A
LECES statement was developed for terrestrial and aquatic ecosystems.
For research evaluating ecological effects, emphasis was placed on recent studies published since
the literature cutoff date of the 2013 Pb ISA that: (1) evaluated effects at concentrations at or near current
environmental concentrations of Pb in soil, water, and sediment and (2) investigated effects on species,
subspecies, or study populations of algae and plants, microbes, invertebrates, or vertebrates at any
lifestage or in any biological community or ecosystem. Exposure concentrations, endpoints, and study
types considered for the 2024 Pb ISA that inform understanding of the ecological effects of Pb in
terrestrial and aquatic systems are summarized further in the LECES statement (Table 12-4). For
exposure concentrations, guidelines were used when screening studies for inclusion. These guidelines
took into consideration data that was current at the time of the 2013 Pb ISA on Pb concentrations in soils,
water, and sediments in the United States (Table 1-1 from the 2013 Pb ISA). The concentration guideline
for literature screening in the 2024 Pb ISA is approximately one order of magnitude higher than upper
bound values from available environmental surveys for soils, water, and sediment (refer to the footnotes
in Table 12-4). For soil, the concentration guideline for screening of terrestrial studies of Pb exposure and
effects was set at approximately 230 mg Pb/kg of soil, although higher concentrations were considered if
the study added new information on a mechanism of action, or if the higher concentration was part of a
series that contributed exposure-response information and included other concentrations below 230 mg
Pb/kg. For aqueous exposures, the concentration guideline for study screening was approximately 10 |ig
Pb/L, although higher concentrations were considered if the study added new information on a
mechanism of action or if the higher concentration was part of a series that contributed exposure-response
information. For sediments, the concentration guideline for study screening was approximately 300 mg
Pb/kg dry weight or lower. Studies at very high concentrations of Pb in soils, water, and sediments were
excluded unless they were part of a series in an experimental exposure-response study and at least one
concentration in the test series was in the ranges stated above (Table 12-4).
In addition to the biological effects described in the LECES statement, other topics within scope
included how chemical and biological modifying factors affect bioavailability in terrestrial, freshwater,
and saltwater environments, as well as studies that address key uncertainties and limitations in the
evidence identified in the 2013 Pb ISA. Site-specific studies in non-U.S. locations that do not contribute
to novel insights into Pb biogeochemistry or effects are considered outside of the scope of the 2024 Pb
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ISA. Studies on mine tailings, biochar, industrial effluent, sewage, ship breaking, bioremediation of
highly contaminated sites, and ingestion of Pb shot, fishing tackle, or pellets are also outside the scope of
the 2024 Pb ISA due to the high concentration of Pb and lack of a connection to an air-related source or
process.
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Table 12-4 Level of Biological Organization, Exposure, Comparison, Endpoint, and Study Design statement to
define the parameters and provide a framework for identifying relevant ecological studies
Level of Biological Organization: Species or subspecies, study populations of vegetation, microbes, invertebrates, or vertebrates, at any
lifestage, or any biological community or ecosystem in terrestrial environments present in the United States or similar to those in the United States.
Exposure: Short or long-term Pb concentrations in exposure media (e.g., soil or diet) that are most relevant to environmental concentrations of Pb
in the United States.3 For soil, the guideline for screening of terrestrial studies of Pb exposure and effects was defined as a concentration of
approximately 230 mg Pb/kg,b with higher concentrations considered if the study elucidates a mechanism or is an acute exposure and at least one
concentration in the test series is in the range described above. Analytically verified exposure concentrations preferred; nominal concentrations
considered in some cases.
Terrestrial
Comparison: A comparison to an unexposed laboratory control, a reference population, or site with no detectable exposure or with lower Pb
exposure.
Endpoint: Species or population effects including effects on growth, reproduction or development, neurobehavioral effects, reduced survival or
fitness, carbon fixation and photosynthesis. At higher levels of biological organization endpoints include changes in community composition,
altered ecosystem processes and functions, such as productivity, community composition, or shifts in genotypes or species, species extirpation,
declines in total number of species or biomass, or decreased species richness.
Study Design: Laboratory, mesocosm, observational or experimental field or gradient studies, or mechanistic modeling studies that estimate the
effect of Pb on an organism, biological population, community, or ecosystem whose processes may be represented quantitatively (e.g., in a
dynamic or steady state).
Level of Biological Organization: Species and subspecies, study populations of vegetation, microbes, invertebrates, or vertebrates, at any
lifestage, or any biological community or ecosystem in freshwater or saltwater environments and transition zones present in the United States, or
similar to those in the United States, excluding the open ocean.
Exposure: Short or long-term Pb concentrations in exposure media (e.g., water, sediment, or diet) that are most relevant to environmental
concentrations of Pb in the United States.3 For freshwater or saltwater, the guideline for screening of Pb exposure and effects was defined as a
concentration of approximately 10 |jg Pb/Lc with higher concentrations considered if the study elucidates a mechanism plausibly relevant at lower
concentrations. For sediments, exposure concentration of approximately 300 mg Pb/kg, dry weight.d For dietary pathways, at least one
experimental group (prey) exposed to approximately 10 |jg Pb/L (aqueous guideline for screening) prior to a feeding study. If a study provides
Aquatic toxicity data on a previously untested organism grouping (such as Class, Order, Family) or for lower concentration studies of an organism with a
protected status, studies were included even if concentrations exceeded the guideline. Analytically verified exposure concentrations preferred;
nominal concentrations considered in some cases.
Comparison: A comparison to an unexposed laboratory control, a reference population, or site with no detectable exposure or with lower Pb
exposure.
Endpoint: Species or population effects including effects on growth, reproduction or development, neurobehavioral effects, reduced survival or
fitness, carbon fixation and photosynthesis. At higher levels of biological organization endpoints include changes in community composition,
altered ecosystem processes and functions, such as productivity, or shifts in genotypes or species, species extirpation, declines in total number of
species or biomass, or decreased species richness.
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Study Design: Laboratory, mesocosm, observational or experimental field or gradient studies or mechanistic modeling studies that estimate the
effect of Pb on an organism, biological population, community, or ecosystem whose processes may be represented quantitatively (e.g., in a
dynamic or steady state).
Pb = lead.
aStudies on mine tailings, industrial effluent, land-applied sewage sludge, ship breaking, bioremediation of highly contaminated sites, and ingestion of Pb shot or pellets are not
within the scope of the ISA due to a high concentration of Pb or lack of a connection to an air-related source or process. Generally excluded are studies of metal mixtures for which
a specific effect of Pb was not separated unless conducted in biological systems with limited experimental evidence. Lastly, most site-specific studies conducted outside of North
America that do not contribute novel insights on Pb biogeochemistry or effects are excluded.
bThe guideline for screening of terrestrial studies of Pb exposure and effects is based on the values reported for soils of the conterminous United States in the 2013 United States
Geological Survey report "Geochemical and mineralogical data for soils of the conterminous United States" (Smith et al.. 2013). This survey was conducted between 2007 and 2013
and sampled three soil horizons (surface, A, and C) at 4,857 nonurban, non-near-road sites. The Q1, median, mean, and Q3 values in surface soil (0-5 cm) for 4841 locations for
which Pb data was available in North American Soil Geochemical Landscapes Project (NASGLP) were 13.5, 18.1, 25.8, and 23.9 mg Pb/kg soil. The Q1, median, mean, and Q3
values in the A horizon (relevant for plants, invertebrates, and microorganisms as well as burrowing mammals and reptiles) for 4,841 locations for which Pb data was available in
NASGLP were 13.2, 17.8, 22.2, and 23.2 mg Pb/kg soil. The 230 mg Pb/kg soil concentration guideline is approximately one order of magnitude higher than the Q3 values from the
survey.
°The guideline for screening of Pb concentration in water is based on United States Geological Survey National Water Quality Assessment sampling for which the 2006 Pb AQCD
reported summary statistics as of the time (U.S. EPA. 2006). The 99th and 95th percentile dissolved Pb values were 5.44 |jg/L and 1.1 |jg/L, respectively (see Table 6-2 in the 2013
Pb ISA) (U.S. EPA. 2013a). A more relevant upper bound value for dissolved Pb would be closer to 1 |jg/L, and 10 |jg/L is one order of magnitude above that value. As dissolved
Pb concentrations in saltwater would be expected to be no higher—and generally, lower—than concentrations in freshwater (due to odds of greater proximity of freshwaters to
anthropogenic sources and less access to mixing), an upper bound for saltwater would reasonably be expected to be lower than that for freshwater concentrations.
dThe guideline for Pb screening in sediment is based on an older survey of urban and reference lake sediments across the United States. (Mahler et al., 2006) and further
supported by evidence from more recent regional survey data. A median 1990s concentration for 35 U.S. sites (Table 2 of (Mahler et al.. 2006)) of 73 mg Pb/kg was reported and
the paper concluded that Pb had decreased since 1970s, with the 1990s median being 40% lower than the 1970s median. For saltwater, Kim et al. (2004) reported samples in a
lower Delaware coastal saltmarsh that would be expected to have much less historic and non-air contamination. The concentrations for the upper depths (0 to 5 cm), dated to reflect
the 90s through the early 2000s, range from 20 to 30 mg/kg. Thus, 30 mg/kg appears to be a more appropriate upper bound value for freshwater and saltwater sediments, and 300
mg Pb/kg is one order of magnitude above that value.
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12.5 Literature Search
The U.S. EPA uses a structured approach to identify relevant studies for consideration and
inclusion in the ISAs. The search for relevant literature in this review began with publishing a Request for
Information FRN (July 7, 2020, 85 FR 40641). This FRN announced the initiation of this Pb NAAQS
review and invited the public to submit relevant research studies and data that have been published,
accepted for publication, or presented at a public scientific meeting since January 1, 2011, providing
overlap with the 2013 Pb ISA wherein the literature considered extended to September 2011. Literature
submitted by the public in response to this FRN can be viewed in the U.S. EPA's HERO database. U.S.
EPA reviewed these studies for relevance following the literature screening process described in this
appendix.
In addition to the Request for Information FRN, the U.S. EPA applied systematic review
methodologies to identify peer-reviewed scientific literature relevant to the 2024 Pb ISA. The literature
searching and screening methodology used for the 2024 Pb ISA generally followed the process depicted
in Figure 12-2. The process began with a combination of keyword searches and citation network searches
to find relevant literature in PubMed and Web of Science published between September 2011 and
December 2020. This literature search strategy was designed to maximize precision2 and recall3 for each
discipline (i.e., health, welfare effects, atmospheric sciences, and exposure). The literature then went
through two levels of screening to identify relevant studies: (1) title and abstract screening using SWIFT-
Active Screener (SWIFT-AS), and (2) full-text screening if the peer-reviewed paper was deemed
potentially relevant after title and abstract screening.
Keyword searches were developed for each appendix using strings of relevant search terms to
capture literature relevant to Pb and the topics in each appendix. For human health search results,
automatic topic classification, a process that uses machine learning to classify references based on a set of
already identified relevant papers, was then used to separate epidemiologic references from experimental
references. In addition to keyword searches, topic-specific citation network searches for all disciplines
were used to identify publications that cite references included in the 2013 Pb ISA. This approach allows
for relevance ranking: given a set of seed references from the 2013 Pb ISA, the more seed references that
a new reference cites, the more likely that new reference is to be relevant. In addition, a small number of
references were also identified for consideration in the 2024 Pb ISA through identification of relevant
literature by U.S. EPA expert scientists; recommendations received in response to the Request for
Information and the Peer Input Workshop; and by review of citations included in previous assessments or
in newly identified literature. Reviewers during the Peer Input Workshop were asked to provide a list of
Precision is the proportion of relevant references relative to all references retrieved in a literature search.
3Recall is the proportion of relevant references identified by screening, relative to the total number of relevant
references that exist.
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additional references (if any) that the U.S. EPA should consider for the ISA, including those published
since the initial literature search.
Following the 2022 Peer Input Workshops and prior to the release of the External Review Dra ft,
the U.S. EPA updated the initial literature searches. These searches were conducted in response to
comments received on the IRP Volume 2 from the CASAC consultation, and feedback received during
the Peer Input Workshops. The updated literature searches targeted key, policy-relevant topics (i.e.,
"scientific information and analyses that address key questions related to the adequacy of the standards"
(U.S. EPA. 2022)) most informative to reviewing the Pb NAAQS to ensure that literature published since
the cutoff date of the initial literature searches was captured. For the selected health effects (i.e., nervous
system, cardiovascular, and reproductive and developmental health effects) the updated literature search
captured epidemiologic and experimental literature published between December 2020 and June 2022.
For effects of Pb in terrestrial and aquatic ecosystems, the updated literature search included the date
range of August 2020 to June 2022 and focused on studies reporting effects on growth, reproduction, and
development or survival. For atmospheric sciences, the same search strings used for the original search
were applied to the date range of August 2020 to June 2022. The U.S. EPA then conducted title and
abstract and full-text screening steps to these additional references.
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Pb = lead.
Figure 12-2 Literature flow diagram for the Pb Integrated Science
Assessment.
12.5.1. Title and Abstract Screening
Consistent with the 2020 Ozone ISA (U.S. EPA. 2020b). the U.S. EPA used SWIFT-AS to
perform the first-level screening of the search results for relevance, based on the title and abstract.
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SWIFT-AS is a web-based literature screening software application that uses machine learning to allow
screeners to efficiently screen literature for relevance (Howard et al.. 2020). It ranks search results by
descending likely relevance using a bag-of-words approach and Latent Dirichlet Allocation, trained by
both the scrccner's inclusion and exclusion decisions and a positive training set, when supplied (Howard
et al.. 2016). The U.S. EPA used such a set of "seed references" (references known to be relevant from
the 2013 Pb ISA). As references are screened and tagged as relevant or not relevant, the ranking model is
further trained to sort the remaining literature, pushing predicted relevant literature to the top of the queue
of references to be screened. U.S. EPA screened literature until SWIFT-AS estimated that 95% of
relevant literature was included, a threshold considered comparable to human error rates (Howard et al..
2020; Cohen et al.. 2006).
12.5.1.1. Atmospheric Science
Initial literature related to air quality, atmospheric chemistry, fate, and transport discussed in
Appendix 1 of the 2024 Pb ISA, Lead Source to Concentration, was identified using a strategy consistent
with the approach described in Volume 2 of the IRP (U.S. EPA, 2022). The search involved both a
citation network search and a keyword search component. For all air sections (Appendix 1, Sections 1.2,
1.3.1, 1.3.4. 1.4. and 1.5). the citation network search identified all publications that cited any references
from the 2013 Pb ISA chapter, Ambient Lead: Source to Concentration (U.S. EPA, 2013a), and a
keyword search was developed to capture additional relevant publications in the Web of Science database
that did not cite any 2013 Pb ISA references. The search string was tested to confirm it would achieve
greater than 99% recall when applied to the 2013 Pb ISA chapter references. Literature for the fate and
transport sections on soil and water (Appendix 1, Sections 1.3.2 and 1.3.3) was obtained in a similar
manner, using the citation network and keyword searches used for terrestrial and aquatic ecosystems
(Section 12.5.1.4). SWIFT-AS was used for title and abstract screening with seed references from the
2013 Pb ISA. Decisions about inclusion or exclusion were guided by the PICOC statement (Table 12-1).
After the Peer Input Workshop (Section 12.7.3), the literature search was updated using the same
two search strings originally applied to the Web of Science database for references published after the
original cutoff date. Consistent with the initial literature search, the U.S. EPA screened these additional
studies for relevance using SWIFT-AS; decisions about relevance were guided by the PICOC statement.
12.5.1.2. Exposure Assessment
Initial literature related to ambient Pb exposure, toxicokinetics, and biomarkers discussed in
Appendix 2 of the 2024 Pb ISA, Exposure, Toxicokinetics, and Biomarkers, was identified using a
keyword search strategy consistent with the approach described in Volume 2 of the IRP (U.S. EPA.
2022). This search involved both a citation network search and a keyword search. The citation network
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search was designed to identify all publications that cited any references from Chapter 3: Exposure,
Toxicokinetics, and Biomarkers of the 2013 Pb ISA (U.S. EPA, 2013a).
Two separate keyword searches were developed to capture additional relevant publications that
did not cite any 2013 Pb ISA references from the Web of Science and PubMed databases, respectively.
The inclusion and exclusion terms used for each search were developed independent of one another to
maximize the relevance for each database. Given the extensive overlap between publications that
contained information on Pb exposure, biomarkers, and toxicokinetics, both keyword searches were
performed on all topics in the appendix. Results from both searches were combined and literature was de-
duplicated.
SWIFT-AS was used for title and abstract screening. The SWIFT-AS algorithm was initially
trained using references from Chapter 3: Exposure, Toxicokinetics, and Biomarkers of the 2013 Pb ISA
(U.S. EPA, 2013a) as seed references. Literature tags were developed to organize results by subsection.
Judgments of inclusion and exclusion were based on guidelines described in the Relevance and Scope
section above (Section 12.4.2).
Following the 2022 Peer Input Workshop, peer input reviewers determined that the U.S. EPA had
identified most of the relevant literature. Suggested additions were screened for relevance and judgments
of inclusion and exclusion were based on guidelines described in the Relevance and Scope section above
(Section 12.4.2).
12.5.1.3. Health
Epidemiologic and experimental studies (i.e., animal toxicology studies) examining health effects
from Pb exposure were targeted using a broad keyword search and citation network search strategy
consistent with Volume 2 of the IRP (U.S. EPA, 2022). U.S. EPA screened the identified literature for
relevance against PECOS statements for each health endpoint (see Section 12.4.3), using SWIFT-AS.
The SWIFT-AS algorithm was trained initially using seed references from the 2013 Pb ISA (U.S. EPA,
2013a).
During this first phase of screening, the U.S. EPA tagged experimental studies reporting health
outcome-related literature that potentially informs the biological or chemical events associated with
phenotypic effects, including in vitro, in vivo (by various routes of exposure), ex vivo, and in silico
studies. Although these studies do not necessarily meet PECOS criteria, they were tracked as a
supplemental evidence stream to inform biological plausibility.
Following the 2022 Peer Input Workshop, the U.S. EPA updated the literature search for the
following health outcome categories using the same keyword and citation network search strategy:
nervous system effects (Appendix 3); cardiovascular effects (Appendix 4); and reproductive and
developmental effects (Appendix 8). The updated literature search focused on key, policy-relevant health
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outcomes for which a substantial body of recent literature conducted at relevant Pb biomarker levels was
expected, as suggested by results from the initial search. Consistent with the initial literature search, the
U.S. EPA screened these additional studies for relevance using SWIFT-AS and the PECOS statements.
12.5.1.4. Welfare—Effects on Terrestrial and Aquatic Ecosystems
Studies potentially relevant to Pb effects in terrestrial or aquatic ecosystems (freshwater and
saltwater) were identified using a broad keyword search and citation network search strategy consistent
with the approach described in Volume 2 of the IRP (U.S. EPA, 2022). The U.S. EPA screened the
identified literature for relevance against LECES statements using SWIFT-AS (Table 12-4). The SWIFT-
AS algorithm was trained initially using seed references from the 2013 Pb ISA (U.S. EPA, 2013a).
Studies that were not within the scope of the ISA or that did not meet the criteria for inclusion based on
title and abstract screening (Section 12.4.4 and Table 12-4) were excluded from further consideration.
Following the 2022 Peer Input Workshop, the U.S. EPA updated the literature search and screened
additional studies in SWIFT-AS for relevance using the LECES statements.
12.6 Study Selection: Full-Text Screening and Evaluation of
Studies
The U.S. EPA performed a second level of screening based on assessment of the full text of the
references remaining after the first-level screening (title and abstract). The U.S. EPA continued to use
relevance criteria outlined in Section 12.4 during full-text screening. Studies selected for inclusion based
on relevance were evaluated for study quality, as described below.
12.6.1. Individual Study Quality
After selecting studies for inclusion based on relevance, individual study quality was evaluated by
considering the design, methods, conduct, and documentation of each study, but not the study results. For
ISAs, the overall individual study quality evaluation process is described in the Preamble (U.S. EPA.
2015b). which outlines a base set of questions for consideration when evaluating the scientific quality of
studies, intended for use in both human health and ecological studies:
• Were the study designs, study groups, methods, data, and results clearly presented in relation to
the study objectives to allow for study evaluation? Were limitations and any underlying
assumptions of the design and other aspects of the study stated?
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• Were the ecosystems, study site(s), study populations, subjects, or organism models adequately
selected, and are they adequately defined to allow for meaningful comparisons between study or
exposure groups?
• Are the air quality, exposure, or dose metrics of adequate quality and are they sufficiently
representative of or pertinent to ambient air?
• Are the welfare effect measurements meaningful, valid, and reliable?
• Were likely covariates or modifying factors adequately controlled or taken into account in the
study design and statistical analysis?
• Do the analytical methods provide adequate sensitivity and precision to support conclusions?
• Were the statistical analyses appropriate, properly performed, and properly interpreted?
Worldwide, formal methods for individual study quality evaluation are much better developed for
human health research than for ecological, atmospheric, and exposure studies. The study quality approach
for health and welfare are described further below. For the 2024 Pb ISA, atmospheric and exposure
studies were considered acceptable if they were published in a peer-reviewed journal, though further
scrutiny was applied during full-text screening of exposure studies to identify whether the exposure
assessment methods were clearly described; the selected exposure assessment methods were appropriate
for the research question evaluated; the assumptions of the method(s) were clearly stated; the
uncertainties and limitations of the methods were clearly stated; and QA testing had been performed. No
studies in the atmospheric or exposure, toxicokinetics, and biomarkers appendices were deemed to have
unacceptable study quality.
Study quality was a final step in full-text screening to decide whether to include a study in the
ISA. Any references that did not pass the study quality review and deemed uninformative for the purposes
of this assessment were excluded from the ISA. Studies that passed both the relevance screening and the
study quality evaluation were included in the ISA. The combination of approaches described in this
section are intended to produce a comprehensive collection of pertinent studies needed to address the key
scientific issues that are examined in the ISA.
12.6.1.1. Health
As described in the Preamble (U.S. EPA. 2015b). causality determinations are informed by
integrating evidence across scientific disciplines (e.g., exposure, animal toxicology, epidemiology) and
related outcomes, and by judgments of the strength of inference in individual studies. For health
outcomes, study quality is evaluated using a uniform approach that considers study strengths and
limitations, including the possible roles of chance, confounding, and other biases that may influence
results. The process for individual study quality evaluation has been refined by discipline with each
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successive ISA based on input and feedback from numerous reviews by CAS AC. Recent IS As have
developed study quality criteria tables to provide clarity on important aspects of study quality for health
outcomes and serve as the foundation for the review of individual health studies (U.S. EPA, 2020b, 2019,
2017, 2016). These aspects describe the characteristics of study elements (e.g., study design, exposure
assessment, potential confounding factors) that can increase or decrease confidence in the study results.
Where possible, study elements, such as exposure assessment and confounding (i.e., bias due to a
relationship with the outcome and correlation with exposures to Pb) are tailored to address factors specific
to health studies of Pb exposure. Thus, judgments on the ability of a study to inform the relationship
between an air pollutant and health vary depending on the specific pollutant being assessed.
Table 12-5 describes the aspects considered in evaluating study quality of animal toxicological
and epidemiologic studies considered for inclusion in the 2024 Pb ISA. The specific aspects of each
domain listed in Table 12-5 are consistent with current best practices for reporting or evaluating health
science data.4 Additionally, the aspects are compatible with published U.S. EPA guidelines related to
cancer, neurotoxicity, reproductive toxicity, and developmental toxicity (U.S. EPA, 2005, 1998, 1996,
1991). These aspects were not used as a checklist to determine if a study should be included or excluded;
the presence or absence of particular features in a study did not necessarily lead to the conclusion that a
study was less informative or should be excluded from consideration in the ISA. Instead, reviewers
considered each element of a study and made a final binary judgment (include or exclude) based on
overall study quality. Study quality considerations for individual studies may be discussed within the
health appendices of the 2024 Pb ISA in instances when specific aspects affect the interpretation of a
study, either increasing or decreasing confidence in study results. Importantly, judgments were made
without considering the outcome of a study (e.g., whether an adverse health outcome was observed), and
these aspects were not used as criteria for determining the causal relationship between Pb exposure and
health effects. As described in the Preamble (U.S. EPA, 2015b), causality determinations were based on
judgments of the overall strengths and limitations of the collective body of available studies and the
coherence of evidence across scientific disciplines. Table 12-5 is not intended to be a complete list of
aspects that define a study's ability to inform the relationship between Pb and health effects, but it
describes the major aspects considered in the 2024 Pb ISA to evaluate studies.
A limited number of studies have been excluded based on consideration of the study quality
aspects described in Table 12-5. For example, specific epidemiologic studies have been excluded due to
the evaluation (solely) of univariate models; lack of statistical power to detect an association; and
inadequate or missing description of methods. In addition, specific toxicological studies were excluded
from consideration because observed effects could not be reliably attributed to Pb exposure; application
of an experimental model that was not intended for use with animals; reporting data that directly conflict
with results of different experiments described in the same publication without explanation, along with
4For example, the National Toxicology Program (NTP) Office of Health Assessment and Translation (OHAT) approach
(Rooney et al., 2014), Integrated Risk Information System (IRIS) Preamble (U.S. EPA, 2013b), ToxRTool (Klimisch et al.,
1997), Strengthening the Reporting of Observational Studies in Epidemiology (STROBE) guidelines (von Elm et al., 2007),
aand Animal Research: Reporting of In Vivo Experiments (ARRIVE) guidelines (Kilkenny et al., 2010).
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mislabeled figures, which together reduce confidence in the conclusions of the study; and for conducting
experiments performed in animals that were not approved by an institutional animal care and use
committee.
To document the study quality evaluation for a subset of the most policy-relevant health studies, a
narrative approach was used to provide nuanced and transparent documentation of the strengths and
limitations that support expert judgment for individual studies. Narrative reviews were completed for
epidemiologic studies of Pb exposure and full-scale IQ in children, which played a significant role in the
development of the Policy Assessment in the 2016 Pb NAAQS review. The study quality tables (Table
12-5) were used to develop prompting questions for each study domain designed to assist in the narrative
documentation of study quality, ensuring the inclusion of consistent information across reviewers. The
narrative reviews, along with the prompting questions, were recorded in HAWC and can be accessed on
the HAWC project page.
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Table 12-5 Scientific considerations for evaluating the strength of inference
from studies on the health effects of Pb
Study Design
Epidemiology
Inference is stronger for studies that clearly describe the primary and any secondary aims of the study, or specific
hypotheses being tested. Information including the age of the population studied, study period, and study location is
used to aid in the interpretation of findings because Pb exposure has declined over time and exposures vary
depending on proximity to Pb sources.
For observational studies of Pb exposure and health outcomes, inference is considered to be stronger for
prospective cohort studies and case control studies nested within a cohort (e.g., for rare diseases) than other case
control, cross sectional, or ecologic studies. Cohort studies can better inform the temporality of exposure and effect.
Other designs can have uncertainty related to the appropriateness of the control group or validity of inference about
individuals from group level data. Study design limitations can bias health effect associations in either direction.
Animal Toxicology
The primary and any secondary objectives of the study, or specific hypotheses being tested should be clearly
described. Studies should include appropriately matched control exposures (e.g., to clean filtered air, time matched).
Studies should use experimental conditions that provoke little concern for uncontrolled variables or different practices
across groups. Groups should be subjected to identical experimental procedures, conditions, and animal care (e.g.,
housing and husbandry).
Study Population/Test Model
Epidemiology
There is greater confidence in results for study populations that are recruited from and representative of the target
population. Studies with high participation and low dropout over time that is not dependent on exposure or health
status are considered to have low potential for selection bias. Clearly specified criteria for including and excluding
subjects, and the reporting of baseline information on participants that are lost to follow up can aid assessment of
selection bias. For populations with an underlying health condition, independent, clinical assessment of the health
condition is valuable, but self-report of physician diagnosis generally is considered to be reliable for respiratory and
cardiovascular diseases.3 Comparisons of groups with and without an underlying health condition are more
informative if groups are from the same source population. Selection bias can influence results in either direction or
may not affect the validity of results but rather reduce the generalizability of findings to the target population.
Animal Toxicology
The animal species and strain used for toxicology investigations must be appropriate for the study goals and have
relevance to a corresponding outcome in humans. Ideally, studies should report species, strain, substrain, genetic
background, age, sex, and weight. Where applicable, approval of study protocols by appropriate institutional animal
care and use committees must be obtained. Unless data indicate otherwise, PECOS-relevant laboratory nonhuman
mammalian species and strains are considered appropriate for evaluating effects of Pb exposure. It is preferred that
the authors test for effects in both sexes across multiple lifestages and report the result for each group separately.
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Pollutant
Epidemiology
The focus is on studies evaluating Pb exposure.
Animal Toxicology
Studies should focus on the effects of Pb exposure on health outcomes; however, information from mixture studies in
which Pb is a component may be informative if the study employs a Pb-only treatment arm with appropriate control
group. Ideally, studies should report the source, purity, and form of Pb (e.g., lead acetate) used.
Exposure Assessment or Assignment
Epidemiology
General population studies using Pb biomarkers (e.g., blood, bone, or tooth Pb concentrations) are emphasized. The
most useful biomarker of exposure is one that reflects the exposure timing and duration that is appropriate to the
underlying pathogenetic processes (e.g., recent, cumulative over lifetime, or cumulative over a developmental^
sensitive window).
Blood Pb concentration (PbB) is typically measured in venous or capillary blood specimens using a variety of
laboratory analytical techniques. Validated analytical methods with lower LODs, such as inductively coupled plasma
mass spectrometry or graphite furnace atomic absorption spectrometry, are preferred. Capillary blood Pb
determinations have greater potential for contamination during collection, resulting in greater measurement error,
particularly at concentrations approaching the LOD. While PbB is most commonly measured in samples of whole
blood, the small fraction of Pb in plasma (<1%) is the more toxicologically active fraction of the circulating Pb.
Bone Pb is most commonly measured in the tibia, calcaneus, patella, or finger bone via X-ray fluorescence. Recent
studies favor measurement of the patella for estimating trabecular bone Pb, because it has more bone mass and
may afford better measurement precision than the calcaneus. Bone measurements are typically expressed in units of
|jg Pb per g bone mineral. This convention may potentially introduce variability into the bone Pb measurements
related to variation in bone density. Notably, lower bone mineral density is associated with greater measurement
uncertainty in bone Pb, which can have important implications for studies in populations for whom low bone mineral
density is more common (e.g., older women).
Measurements of Pb in hair, saliva, nails, urine, and feces suffer from high interlaboratory variability, low
reproducibility, and a lack of reliable reference values. A more detailed discussion of exposure biomarkers can be
found in Appendix 2.
Animal Toxicology
For this assessment, the administration of Pb by oral, inhalation, or intravenous routes are considered relevant.
Studies that resulted in measured blood Pb levels <30 |ig/dl_ will be used in the health section narratives.15 Studies
should characterize Pb concentration, environmental temperature and relative humidity, and/or have measures in
place to adequately control the exposure conditions. All studies should include exposure control groups (e.g., dosing
vehicle, or no Pb treatment) that are appropriate to the route, duration of exposure, and study design. Studies should
randomize assignment to exposure groups and, where possible, conceal allocation to research personnel. Blinding of
research personnel to study group may not be possible due to animal welfare and experimental considerations;
however, differences in the monitoring or handling of animals in all groups by research personnel should be
minimized.
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Outcome Assessment
Epidemiology
Inference is stronger when outcomes are assessed or reported without knowledge of exposure status. Knowledge of
exposure status could produce artifactual associations. Confidence is greater when outcomes assessed by interview,
self-report, clinical examination, or analysis of biological indicators are defined by consistent criteria and collected by
validated, reliable methods. Independent, clinical assessment is valuable for incidence of disease, but report of
physician diagnosis has shown good reliability. Validated questionnaires for subjective outcomes such as symptoms
are regarded to be reliable,0 particularly when collected frequently and not subject to long recall. For biological
samples, the stability of the compound of interest and the sensitivity and precision of the analytical method is
considered. If not based on knowledge of exposure status, errors in outcome assessment tend to bias results toward
the null.
Animal Toxicology
Endpoints should be assessed in the same manner for control and exposure groups (e.g., time after exposure,
evaluation methods/procedures, endpoint evaluation) using valid, reliable methods. Wherever possible, the limit of
detection for quantitative assays should be given. For each experiment and each experimental group, including
controls, precise details of all procedures carried out should be provided. Time of the endpoint evaluations is a key
consideration that will vary depending on endpoint evaluated. Endpoints should be assessed at time points that are
appropriate for the research questions. Additionally, in order to preclude reporting bias, studies should report results
for all experimental procedures conducted. All animals used in a study should be accounted for, and rationale for
exclusion of animals (e.g., attrition) or data should be specified and reasonable given the study design.
Other Potential Confounding Factorsd
Epidemiology
Factors are considered to be potential confounders if demonstrated in the scientific literature to be related to health
effects and correlated with Pb. Not accounting for confounders can produce artifactual associations; thus, studies
that statistically adjust for multiple factors or control for them in the study design are emphasized. Less weight is
placed on studies that adjust for factors that mediate the relationship between Pb and health effects, which can bias
results toward the null. Confounders vary according to study design and health effect of interest, and may include,
but are not limited to the following: socioeconomic status, parental caregiving, race (as a proxy measure for a
complex set of social factors), age, medication use, smoking status, noise, urbanicity, and environmental and/or
occupational exposures.
Animal Toxicology
Preference is given to studies using experimental and control groups that are matched for individual level
characteristics (e.g., strain, sex, body weight, litter size, and food and water consumption) and time varying factors
(e.g., seasonal and diurnal patterns).
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Statistical Methodology
Epidemiology
Multivariable regression models that include potential confounding factors are emphasized. Studies of pollutant
mixtures can be informative if health effects of exposure to Pb, presumably a component of the mixture, are also
examined separately. Such studies can provide insight into potential modification of the criteria pollutant's effect by
other individual pollutants or by a broader pollutant mixture. Models with interaction terms aid in the evaluation of
potential confounding as well as effect modification. Sensitivity analyses with alternate specifications for potential
confounding inform the stability of findings and aid in judgments of the strength of inference from results. In the case
of multiple comparisons, consistency in the pattern of association can increase confidence that associations were not
found by chance alone. Statistical methods that are appropriate for the power of the study carry greater weight. For
example, categorical analyses with small sample sizes can be prone to bias results toward or away from the null.
Statistical tests such as correlation coefficients, t tests, and chi-squared tests are not considered sensitive enough for
adequate inferences regarding Pb health effect associations. For all methods, the effect estimate and precision of the
estimate (i.e., width of 95% CI) are important considerations ratherthan statistical significance.
Animal Toxicology
Statistical methods should be clearly described and appropriate for the study design and research question (e.g.,
correction for multiple comparisons). Specific sample sizes are not criteria for inclusion or exclusion; ideally, the
sample size should provide adequate power to detect hypothesized effects. Because statistical tests have limitations,
consideration is given to both trends in data and reproducibility of results. Results should be presented quantitatively
in the appropriate format for the data (e.g., continuous data ideally should not be presented as categorical or
dichotomized) and separately by sex and cohort.
CI = confidence interval; LOD = limit of detection; Pb = lead; PbB = blood lead concentra
aMurqia et al. (2014); Weakley et al. (2013); Yang et al. (2011); Heckbert et al. (2004); Barr et al. (2002); Muhaiarine et al. (1997);
Toren et al. (1993).
bStudies not including a blood lead biomarker were tracked during study screening but were not included/evaluated in the health
section narratives.
cBurney et al. (1989).
dMany factors evaluated as potential confounders can be effect measure modifiers (e.g., season, comorbid health condition) or
mediators of health effects related to Pb (e.g., comorbid health condition).
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12.6.1.2. Welfare—Effects on Terrestrial and Aquatic Ecosystems
Generally, the field of study quality evaluation is much more robust for human health research
than for ecological research. However, study quality is still very important for ecological research, and
U.S. EPA staff have relied on the criteria listed in the Preamble as criteria for reviewing the quality of
individual studies within the 2024 Pb ISA. A limited number of studies were excluded based on
consideration of these study quality questions and application of the LECES statement. The main reasons
studies were eliminated: exposure concentrations that exceeded concentration guidelines, as specified in
the LECES (Table 12-4); no report of Pb concentration; Pb was part of a mixture of metals with no testing
of the independent effect of Pb; a lack of statistical testing for endpoints of interest; inadequate or missing
description of methods; or inadequate study design.
12.7 Peer Review and Public Participation
Peer review is an important component of any scientific assessment, as formalized in the
guidance found in the U.S. EPA's Peer Review Handbook (U.S. EPA. 2015a). The 2024 Pb ISA follows
the policies and procedures identified therein. Additionally, the 2024 Pb ISA is designated as a Highly
Influential Scientific Assessment, which is defined by the Office of Management and Budget's Final
Information Quality Bulletin for Peer Review (hereafter, "Peer Review Bulletin") as:
A subset of Influential Scientific Information that is a scientific assessment (i.e., an evaluation of a
body of scientific or technical knowledge, which typically synthesizes multiple factual inputs,
data, models, and assumptions and applies the best professional judgment to bridge uncertainties
in the available information) that "could have a potential impact of more than $500 million in any
year on either the public or private sector" or "is novel, controversial, or precedent-setting, or has
significant interagency interest."
(https://obamawhitehouse.archives.gov/omb/memoranda fV2005 m05-03A.
As such, there are additional review and transparency steps required in the release of this information
(e.g., public comment). These review and public participation steps are described in the subsequent
sections.
12.7.1. Request for Information
Consistent with the Preamble (U.S. EPA. 2015b). a Request for Information was published in the
Federal Register on July 7, 2020 (85 FR 40641). The purpose of this Request for Information was
announcing the beginning of the review cycle of the air quality criteria and the Pb NAAQS and inviting
the public to submit relevant research studies and data that had been published, accepted for publication,
or presented at a public scientific meeting since January 1, 2011. The public was given 60 days to respond
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to this FRN; the U.S. EPA received eight comments via the Federal eRulemaking Portal
(http ://www. regulations .gov. Docket ID: EPA-HQ-OAR-2020-0312). Literature submitted by the public
in response to this FRN can be viewed in the U.S. EPA's HERO database.
12.7.2. Integrated Review Plan
Following the Request for Information, the U.S. EPA prepared a multi-volume IRP: Volume 1
provides background information on the air quality criteria and standards for Pb; Volume 2 addresses the
general approach for the review and planning of the ISA; and Volume 3 is the planning document for
quantitative analyses considered in the policy assessment. Volume 2 of the IRP (U.S. EPA. 2022). which
describes the plan for developing the ISA, was discussed by CASAC at a public meeting on April 8.
2022. Availability of Volume 2 of the IRP for public comment was announced in the Federal Register on
March 10, 2022 (87 FR 13732). The public was given the opportunity to respond, and the U.S. EPA
received one public comment via the Federal eRulemaking Portal (http://www.regulations.gov. Docket
ID: EPA-HQ-OAR-2020-0312-0010).
Following the April CASAC public meeting, documentation of the meeting and written
comments from individual CASAC members were sent to the U.S. EPA Administrator in a letter dated
April 22, 2022 (https://casac.epa.gov/ords/sab/f?p=l 13:12:17516491975646::: 12).
12.7.3. Peer Input
The role of peer input is described in the Preamble, as well as the Peer Review Handbook (U.S.
EPA. 2015a. b). After a thorough literature search and screening process, the U.S. EPA developed
preliminary draft appendices for initial peer input. Causality determinations had yet to be developed. Peer
input is a process that allows the U.S. EPA to gather early-in-the-process feedback from subject-matter
experts, internal and external to the U.S. EPA, to ensure that the ISA captures relevant new literature and
is focused on the most policy-relevant findings. Peer input serves as a supplement to other peer-review
mechanisms and does not replace a thorough external peer review by CASAC.
Peer input for the 2024 Pb ISA occurred as a series of four webinar workshops, which the U.S.
EPA announced in an FRN on May 6, 2022 (87 FR 27147, Docket ID: EPA-HQ-ORD-2020-0701). The
four workshops were organized by subject: Effects of Pb in Terrestrial and Aquatic Ecosystems;
Epidemiologic and Toxicological Evidence for Health Effects of Pb Exposure; Ambient Pb: Source to
Concentration; and Exposure, Toxicokinetic, and Pb Biomarkers. Workshops were facilitated by U.S.
EPA's contractor, ICF. Peer input reviewers were selected by ICF, with input from U.S. EPA, in
accordance with U.S. EPA's Peer Review Handbook (U.S. EPA. 2015a).
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Peer input reviewers were given the following charges:
• Correct technical errors and identify critical gaps.
• Consider how clearly and logically the appendices and content within the sections are organized.
• Indicate how accurately scientific information is characterized, whether advances in knowledge in
the recent literature have been adequately highlighted, and whether emphasis has been placed on
the most informative, policy-relevant literature.
• Identify any key studies missing, (including those published after the early 2021 literature search
dates for the draft materials), especially any associated with the effects of Pb from ambient air.
Provide full citations for suggested references.
• Indicate any specific issues that should be considered or highlighted that will be important for
integrating evidence across disciplines.
There were additional topic-specific charge questions. Peer input reviewers were not asked to correct
typos or grammatical errors.
During the workshops, peer input reviewers affirmed that the U.S. EPA included the relevant
literature, though some additional studies were identified for U.S. EPA's consideration. Following the
workshop, the U.S. EPA considered comments and incorporated revisions based on the reviewers"
feedback. Suggested studies were screened for relevance as described for the initial literature searches and
incorporated if they met the inclusion criteria (see Sections 12.4 and 12.5.1).
12.7.4. Internal Technical Review and Clearance
The U.S. EPA guidelines, such as the U.S. EPA's Peer Review Handbook (U.S. EPA. 2015a).
recommend an internal technical review process prior to any external dissemination of scientific
information. Consistent with this guidance, the draft ISA was reviewed by U.S. EPA subject-matter
experts. Following the technical review, the U.S. EPA revised the document based on the reviewers"
comments prior to submitting this document for formal U.S. EPA clearance. This final document was
cleared for public release following clearance policy and procedures.
12.7.5. Clean Air Scientific Advisory Committee Peer Review
CASAC served as the official peer review mechanism for the 2024 Pb ISA. Two sections of the
Clean Air Act, Sections 108 and 109 [42 U.S.C. 7408 and 7409], govern the periodic review and
establishment of the NAAQS (U.S. EPA. 2020a). With respect to CASAC, Section 109(d)(2) addresses
the appointment and advisory functions of an independent scientific review committee. Section
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109(d)(2)(A) requires the Administrator to appoint this committee, which is to be composed of "seven
members including at least one member of the National Academy of Sciences, one physician, and one
person representing State air pollution control agencies." Section 109(d)(2)(B) states that the independent
scientific review committee periodically "shall complete a review of the criteria... and the national
primary and secondary ambient air quality standards... and shall recommend to the Administrator any
new... standards and revisions of existing criteria and standards as may be appropriate..." Since the early
1980s, this independent review function has been performed by CASAC. More information on the
CASAC peer review can be found on the U.S. EPA CASAC website.
The draft ISA was released for public comment and CASAC review on March 31, 2023. It was
posted on U.S. EPA's website and its availability was announced in a Federal Register Notice (88 FR
19302). CASAC met on June 13-14. 2023. to review the draft ISA, and met virtually on August 23-24.
2023. to further discuss its consensus comments. The final CASAC letter was sent to the U.S. EPA
Administrator on September 18, 2023. The letter conveyed the committee's consensus advice as well as
comments from individual committee members. The Administrator responded to CASACs letter on the
draft ISA on October 30, 2023. CASAC's letter and U.S. EPA's response are viewable on CASAC's
website.
CASAC, in its letter, found the ISA "to be a comprehensive assessment of the available science
relevant to understanding the health and welfare effects of lead (Pb)" (EPA-CASAC-23-003;
https://casac.epa.g0v/0rds/sab/r/sab apex/casac/activitv?p 18 id=2637&clear=l8&session=22281838571
132#report). CASAC shared recommendations for strengthening and improving the ISA and noted that
"with these recommended changes, the [ISA] will serve as a scientifically-sound foundation for the
agency's review of the National Ambient Air Quality Standards (NAAQS) for Pb." The CASAC
recommended clarifying text to strengthen the characterization of health and exposure evidence as well as
revisions to the causality determinations for some health outcomes. In response to that advice, the U.S.
EPA revisited the evidence supporting the draft determinations for several outcomes, including adult
cognitive function, pregnancy and birth outcomes, female reproductive function, immune system effects,
and mortality. In addition, across the exposure and health appendices, the U.S. EPA clarified the lines of
evidence and rationales that support causality determinations; bolstered the ISA's characterization of the
strengths and limitations of various study designs, analytic approaches, and exposure biomarkers;
expanded discussion of the relationship of particle size to exposure pathways, absorption of Pb into the
blood, and models of Pb exposure-blood Pb relationships; and improved consistency in how the ISA
characterizes study results.
CASAC expressed general agreement with the U.S. EPA's characterization of the sources, fate
and transport, and measurement of Pb, as well as its effects on terrestrial and aquatic biota. In response to
CASAC's advice on Appendix 1 (LeadSource to Concentration), the U.S. EPA included additional
figures on emission and concentration trends; added more detail on aviation gas and fire emissions;
expanded the discussion of monitoring requirements and the Pb monitoring network; and recognized
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monitoring needs and interventions related to high daily concentrations identified by CASAC. In response
to CASAC's advice on ecological effects, the U.S. EPA provided additional clarification on a few specific
topics as indicated, edited other content for brevity, and revisited the evidence supporting the draft
causality determination for neurobehavioral effects in freshwater invertebrates.
12.8 Quality Assurance and Quality Control
QA helps ensure that the U.S. EPA conducts high-quality science that can be used to inform
policymakers, industry, and the public. Agency-wide, the U.S. EPA Quality Program provides the
framework for planning, implementing, documenting, and assessing work performed by the Agency, and
for carrying out required quality assurance and quality control (QA/QC) activities. Additionally, the
Quality Program covers the implementation of the U.S. EPA Information Quality Guidelines and the U.S.
EPA Environmental Information Quality Policy and Procedures (U.S. EPA. 2023a. b, 2002). The 2024 Pb
ISA follows all Agency guidelines to ensure a high-quality document.
For the ISA Program, management of quality assurance is documented in a Program Quality
Assurance Project Plan (PQAPP), which describes the technical approach and QA/QC procedures
associated with the ISA Program. QA objectives and measurement criteria detailed in the PQAPP have
been employed in developing the 2024 Pb ISA. QC checks were conducted on numerical entries
throughout the ISA. At a minimum, numerical values from every fifth citation were verified for accuracy
by an independent U.S. EPA staff member against the original source, and any errors were subsequently
corrected. Furthermore, the U.S. EPA's HERO database has its own QC processes, as documented in
HERO's Quality Assurance Project Plan (QAPP).
The 2024 Pb ISA is classified as QA Category A, which requires at least one audit to be
completed. During assessment development, the Pb ISA underwent two Technical System Audits in July
2022 and July 2023 by an independent contractor, Neptune and Company, Inc. The auditor identified no
major findings and verified that QC procedures were adequately performed and documented in
accordance with QA procedures.
12.9 Conclusion
This appendix describes the overall process of developing the Pb ISA: literature search and
screening methods; study quality evaluation; peer input and peer review; documentation; and QA.
Overall, the U.S. EPA has a robust set of policies and procedures in place to ensure the highest quality
products. In developing the 2024 Pb ISA, the U.S. EPA has followed all the appropriate processes and
endeavored to add additional steps as practicable and needed (e.g., use of SWIFT-AS, scoping statements,
and documentation of individual study quality).
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