United States
Environmental Protection
»m Agency
EPA/600/R-23/061
March 2023
www.epa.gov/isa
Integrated Science
Assessment for Lead
Appendix 2: Exposure, Toxicokinetics, and
Biomarkers
External Review Draft
March 2023
Health and Environmental Effects Assessment Division
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
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DISCLAIMER
1 This document is an external review draft for peer review purposes only. This information is
2 distributed solely for the purpose of predissemination peer review under applicable information quality
3 guidelines. It has not been formally disseminated by the Environmental Protection Agency. It does not
4 represent and should not be construed to represent any agency determination or policy. Mention of trade
5 names or commercial products does not constitute endorsement or recommendation for use.
6
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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://cfpub.cpa.gov/ncca/isa/rccordisplav.cfm ?dcid=357282.
Front Matter
Executive Summary
Integrative 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
LIST OF TABLES 2-v
LIST OF FIGURES 2-vi
ACRONYMS AND ABBREVIATIONS 2-viii
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 2-54
2.2.3 Elimination 2-63
2.3 Pb Biomarkers 2-65
2.3.1 Bone-Pb Measurements 2-65
2.3.2 Blood-Pb Measurements 2-66
2.3.3 Urine-Pb Measurements 2-67
2.3.4 Pb in Other Biomarkers 2-67
2.3.5 Relationship between Pb in Blood and Pb in Bone 2-69
2.3.6 Relationship between Pb in Blood and Pb in Soft Tissues 2-79
2.4 Studies of Pb Biomarker Levels 2-83
2.4.1 Pb in Blood 2-83
2.4.2 Pb in Bone 2-94
2.4.3 Pb in Urine 2-95
2.4.4 Pb in Other Biomarkers 2-97
2.5 Empirical Models of Pb Exposure-Blood Pb Relationships 2-98
2.5.1 Air Pb-Blood Pb Relationships in Children 2-99
2.5.2 Air Pb-Blood Pb Relationships in Adults 2-108
2.5.3 Soil Pb-Blood Pb Relationships 2-109
2.6 Biokinetic Models of Pb Exposure-Blood Pb Relationships 2-112
2.7 Summary and Conclusions 2-114
2.7.1 Exposure 2-115
2.7.2 Toxicokinetics 2-116
2.7.3 Pb Biomarkers 2-117
2.7.4 Air Pb-Blood Pb Relationships 2-119
2.8 References 2-120
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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-PM2.5 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 FDA Total Diet Study (2014-2016) and
What We Eat in America (2009-2014) food consumption data. 2-18
Table 2-5 Contribution of maternal blood Pb to breast milk at 1-3 months
postpartum. 2-23
Table 2-6 Pb content in various consumer products. 2-25
Table 2-7 Co-contaminants in Pb sources. 2-31
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, NHANES 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. 2-51
Table 2-11 Blood-Pb concentrations in the U.S. population. 2-83
Table 2-12 Urine-Pb concentrations in the U.S. population. 2-95
Table 2-13 Summary of estimated slopes for blood Pb-to-air Pb slope factors in
children. 2-101
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LIST OF FIGURES
Figure 2-1 Conceptual model of multimedia Pb exposure. 2-3
Figure 2-2 Distribution of Pb in road dust samples collected in three industrial and
mining towns located in southern Poland. 2-54
Figure 2-3 Plot of blood and plasma Pb concentrations measured in adults and
children. 2-57
Figure 2-4 Relationship between Pb intake and blood Pb concentration in infants
(n = 105, age 13 weeks, formula fed). 2-57
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. 2-59
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. 2-73
Figure 2-7 Half-times of Pb in blood as reported by Specht et al. (2019b). 2-74
Figure 2-8 Simulation of relationship between blood Pb concentration, bone Pb, and
body burden in adults. 2-77
Figure 2-9 Simulation of blood and soft tissue (including brain) Pb in children and
adults who experience a period of increased Pb intake. 2-80
Figure 2-10 Simulation of blood and brain Pb in children and adults who experience a
period of increased Pb intake. 2-81
Figure 2-11 Relationship between Pb in urine, plasma, blood, and bone. 2-82
Figure 2-12 Temporal trend in blood Pb concentrations. 2-85
Figure 2-13 Blood Pb cohort means versus year of exam. 2-86
Figure 2-14 Blood Pb geometric means versus year of NHANES exam by
race/ethnicity. 2-87
Figure 2-15 Geometric mean childhood blood Pb levels assessed between 1 and 8
years old, stratified by race/ethnicity. 2-89
Figure 2-16 Slope factors for blood Pb as a function of air Pb. 2-105
Figure 2-17 Blood Pb versus soil Pb for two New Orleans surveys completed in 2001
and 2017. 2-110
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Figure 2-18 Comparison of slope factors in New Orleans data on linear-linear (top)
and log-log (bottom) plots. 2-111
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ACRONYMS AND ABBREVIATIONS
(i-SRXRF microbeam synchrotron radiation X-ray
fluorescence
1OOLL 100 octane, low lead
AALM All-Ages Lead Model
AAS atomic absorption spectroscopy
AERMOD American Meteorological
Society/Environmental Protection
Agency Regulatory Model
AF absorbed fraction
AHHS American Healthy Homes Survey
ALAD S-aminolevulinic acid dehydratase
ALF artificial lysosomal fluid
AQCD Air Quality Criteria Document
AQS Air Quality System
ASV anodic stripping voltammetry
BLL blood lead level
BLRV blood lead reference value
BMI body mass index
Cal EPA California Environmental Protection
Agency
CDC Centers for Disease Control and
Prevention
CR creatine
CSMR chloride to sulfate mass ratio
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
GFR glomerular filtration rate
GI gastrointestinal
GM geometric mean
GSD geometric standard deviation
HA Housing Authority
HC hydrocarbons
HFE hemochromatosis
ICP-AES inductively coupled plasma atomic
emission spectroscopy
ICP-MS inductively coupled plasma mass
spectrometry
ICRP International Commission on
Radiological Protection
IDF Israeli Defense Force
IEUBK Integrated Exposure Uptake Biokinetic
IMPROVE Interagency Monitoring of Protected
Visual Environments
IRLs interim reference levels
ISA Integrated Science Assessment
IVBA in vitro bioaccessibility
KXRF K-shell X-ray fluorescence
LA-ICP-MS laser ablation inductively coupled
plasma mass spectrometry
LOD limit of detection
LSL lead service line
LTC loading to concentration
MG Mahayogaraj Guggulu
MMB multimedia biomarker
mo month, months
NAAQS National Ambient Air Quality
Standards
NATA National Air Toxics Assessment
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-to-income ratio
PM particulate matter
PUFA polyunsaturated fatty acids
RBA relative bioavailability
RBCs red blood cells
RSD relative standard deviation
SAB Scientific Advisory Board
SD standard deviation
SES socioeconomic status
SHEDS Stochastic Human Exposure and Dose
Simulation
SLL soil lead level
TDS Total Diet Study
TRI Toxics Release Inventory
TSP total suspended particles
VA Veterans Affairs
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WQS weighted quantile sum XRF X-ray fluorescence
WWEIA What We Eat in America yr year, years
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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 onwards. 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 overtime. 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 (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://cfpub.epa.gov/ncea/isa/recordisplav.cfm?deid=357282).
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 j^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 Venn diagram (Figure 2-1) below depicts the various
pathways that Pb can take through the environment to reach a human being.
The "air/soil/water" arrows in the diagram show humans, plants, and animals can be exposed to
Pb through contact with Pb-containing media. These exposures are considered air-related if Pb passed
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 also
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.
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Figure 2-1 Conceptual model of multimedia Pb exposure.
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. 2006a) 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://cfpub.epa.gov/ncea/isa/recordisplav.cfm?deid=357282). 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 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) or inductively coupled plasma mass
spectrometry (ICP-MS) (Yu et al.. 2022). It has also been measured by EPA method 6200, which uses
portable X-ray fluorescence (XRF) (Obeng-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 primary Pb source causing an individual's BLL to become elevated. The
ratio of three isotopes (206Pb, 207Pb, and 208Pb) 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). Becker et al. (2022) used isotope ratios (208Pb/206Pb and 206Pb/207Pb) 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, indoor dust, and paint chips. This suggests that
another unsampled Pb source in the home or elsewhere was causing this child's EBLL.
Xue etal. (2022) developed a generalizable approach to identify locations of hotspots for Pb
exposure based on childrens' 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
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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 2014-2016. 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 (Cimorclli 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).
Biokinetic models have been developed at 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
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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. 2006a). 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 an 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 IEUBKto 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 blood lead level
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. 2013V
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 AQCD (U.S. EPA. 2006a)
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 particulate matter (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 . D. Outdoor
Study Location Metrjc pJlQ^ pb Indoor Pb pb
Clayton, 1999,
IL, IN, Ml,
Med.
July
13
6.6
8.5
14003@@author-year}
MN, OH, Wl
Pb-PMso
1995-
(ng/m3)
May 1997
Adgate, 2007,
Minneapolis-
Avg.
Spring,
6.2
3.4
2.0
156196@@author-year}
St. Paul, MN
Pb-PM2.5
Summer,
(ng/m3)
Fall 1999
Molnar, 2007,
Stockholm,
Avg.
December
Homes: 3.4
Homes: 4.5
156774@@author-year}
Sweden
Pb-PM2.5
2003-July
Schools:
Schools:
(ng/m3)
2004
2.5
4.6
Preschools:
Preschools:
1.8
2.6
Tovalin-Ahumada, 2007,
Mexico City,
Med.
April-May
26
56
190165@@author-year}
Mexico
Pb-PM2.5
2002
(ng/m3)
Puebla,
Med.
April-May
4
4
Mexico
Pb-PM2.5
2002
(ng/m3)
Pekey, 2010,
Kocaeli,
Avg.
May-June
Summer:
Summer:
667712@@author-year}
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, 2007,
Windsor,
Med.
April 2004
311
124
221
91084@@author-year}
Ontario,
Pb-PMzs
Canada
(mg/kg)
Pb = lead; PM = particulate matter
1 Past studies have shown personal Pb-PM concentrations to be higher than indoor or outdoor
2 concentrations. The National Human Exposure Assessment Survey (NHEXAS) study (Clayton et al..
3 1999) cited in the 2006 Pb AQCD (U.S. EPA. 2006a). which sampled Pb in multiple exposure media
4 across six states in EPA Region 5, found personal air Pb concentrations to be significantly higher than
5 indoor or outdoor Pb concentrations. Adgate et al. (2007) found average personal Pb-PM2 5 concentrations
6 to be roughly three times higher than outdoor Pb-PM2 5 concentrations and roughly two times higher than
7 indoor Pb-PM2 5 concentrations.
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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-PM2 5
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 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-PM2 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, 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
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Site3
Season
Outdoor - Mean (ng/m3)
Indoor - Mean
(ng/m3)
Personal - Mean
(ng/m3)
5
Summer
15.0
12.0
11.0
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 sourced 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 PMio-2.5to
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 etal.. 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://cfpub.cpa.gov/ncca/isa/rccordisplav.cfm'Mcid=357282) 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 aviation gas can also contribute to Pb concentrations at and
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around airports, as discussed in Section 1.2.1 of this ISA
(https://cfpub.cpa.gov/ncca/isa/rccordisplav.cfm?dcid=357282). 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://cfpub.epa.gov/ncea/isa/recordisplav.cfm?deid=357282).
Airborne Pb-PM from motor vehicle combustion tends to be smaller than that emitted by
industrial sources, resuspended soil, and tire/brake wear (U.S. EPA. 2013; Schauer et al.. 2006). 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 aviation 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) aviation fuel 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://cfpub.epa.gov/ncea/isa/recordisplav.cfm?deid=357282). 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
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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. 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://cfpub.cpa.gov/ncca/isa/rccordisplav.cfm?dcid=357282) 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 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). A study of 170 homes in Appleton, WI found a range of Pb concentrations
between 47 and 32,483 |ag 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
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. (2010), and (Mielke et al.. 2016).
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 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.. 201 lb) 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. (Miclkc 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 (Miclkc 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. (201 la) 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 etal. (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 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). 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.
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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.
Ozkavnak 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 Lownev (2012) suggests particle adherence occurs below 150 |_im. Finer soil particles
(<63 (.un 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. 2006a) 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 (Zotaetal.. 2011) to 470 mg/kg (Spalinger et al.. 2007). although locations
sampled within buildings (e.g., floors, windowsills) and sampling procedures varied by study.
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 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 (Zotaetal.. 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
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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 (Miclkc 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 jj.g/ft2 or a windowsill dust Pb level >100 jj.g/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 jj.g/ft2, 13.9 jj.g/ft2,
and 31.2 jj.g/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 refinishing
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,
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). 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..
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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 jj.g/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 values (0.578) that were overall most similar to the evaluation dataset slope (0.440), y-intercept
(5.511), and R2 values (0.473) 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 (Flannerv and Middleton. 2022).
The 2006 Pb AQCD (U.S. EPA. 2006a) 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. 2006a). 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 j^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
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-
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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),
female children 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.
Table 2-4 Dietary exposures to Pb based on FDA Total Diet Study (2014-2016)
and What We Eat in America (2009-2014) food consumption data.
Dietary Pb Exposure (|jg/day)
Mean 90th Percentile
Reference Population BoTrTi BouZ* HybridC BouTi BoX HybridC
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
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Dietary Pb Exposure (|jg/day)
Mean 90th Percentile
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
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 x LOD.
Both Gavelek et al. (2020) and Spungen (2019) limited the scope of their analyses to data
collected from 2014 onwards because FDA started using ICP-MS in 2014 to measure elemental
concentrations, as opposed to atomic absorption spectroscopy (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
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
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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
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,
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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 j^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 j^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.
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
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minerals at a neutral pH (Renner. 2006; Vasquez et al.. 2006). Because water quality depends on many
factors, 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. 2003b).
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 217-13,200 (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
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 j^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
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pipes or galvanic corrosion of partially replaced LSLs. Pb concentrations in drinking water for these
homes did not fall below 10 j^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 (Triantafvllidou 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. 2006a). 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 j^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 etal.. 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 etal. (2019)). These factors likely predominately reflect effects on
maternal blood Pb and are not further considered here.
Table 2-5
Contribution of maternal blood Pb to breast milk at 1-3 months
postpartum.
Milk Pb/
Blood Pb
Milk Pba Blood Pba
(pg/L) (Mg/dL)
nb
Location
Sample Yr
Reference
0.10
0.8 ±0.7 7.7 ±4.0
81
Mexico City, Mexico
Ettinqer et al. (2014)
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Milk Pb/
Blood Pb
Milk Pba
(M9/L)
Blood Pba
(H9/dL)
nb
Location
Sample Yr
Reference
1997-1999
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 etal. (1984)
0.25
0.73 ± 0.70
29 ± 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
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, r2 = 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 (201IV 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 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 (r2 = 0.72, p < 0.01) and mothers (r2 = 0.62,
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1 p < 0.01). Thus, the contribution of breast milk Pb itself to infants' blood Pb may be overestimated by
2 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
3 Consumer products have been identified as a source of Pb exposure in previous reviews (U.S.
4 EPA. 2013. 2006a'). This subsection builds upon discussions from the 2013 Pb ISA (U.S. EPA. 2013).
5 highlighting consumer products found in the recent literature, detailing their corresponding Pb content,
6 and further summarizing trends since that ISA (U.S. EPA. 2013).
7 Table 2-6 shows Pb content found in several consumer products, including spices, traditional
8 medicines, cosmetics, toys/baby products, pottery, and tobacco. This table is an update of Table 3-7 in the
9 2013 Pb ISA (U.S. EPA. 2013V
10 Although products purchased in the United States have been found with detectable Pb content,
11 most consumer products identified with detectable Pb content originate in countries abroad (Table 2-6). In
12 many cases, these products were purchased outside the United States and brought into the country by the
13 consumer. For example, in an analysis of 1,496 of the consumer products sampled by the New York City
14 Department of Health and Mental Hygiene (DOHMH) between 2008 and 2017, Hore et al. (2019)
15 reported 45% of spices purchased abroad had Pb content above 2 ppm, as opposed to 13% of sampled
16 spices purchased in the United States. The 2-ppm threshold is the permissible limit in food additives that
17 is used by DOHMH as a guidance limit (Hore et al.. 2019).
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)
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)
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Product
Category
Product
Location of
Purchase
Country of
Origin
Pb Content (units)
Reference
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 pg/cigarette
Afridietal. (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)
Daw Tway
Not specified:
Myanmar or
United States
Myanmar
Median: 520 pg/g
Ritchev et al. (2011)
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Product Product Location of Country of Pb Content (units) Reference
Category Purchase Origin
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; 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 etal.. 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 j^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. Pb exposure occurs through
inhalation or ingestion of Pb in dust and skin contact with leaded bullets. Numerous studies have shown
the connection between EBLLs and ammunition, particularly in indoor and outdoor shooting ranges. 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). A few recent studies examining the relationship between shooting ranges
and Pb exposure are highlighted below.
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% afterthe basic
training course to 89% afterthe 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.
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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 jj.g/ft2 but was reduced to 14 jj.g/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 2010-2011 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 j^ig/dL. 85% of vehicle dust samples and 49% of house dust samples were >
40 jj.g/ft2. Children's BLLs decreased 9.9 (ig/dL on average after 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 j^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).
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 3 categories were found to be
associated with Pb in house dust, pointing to overlapping vulnerabilities. Pb in homes" dust ranged from
20-8,310 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
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25 (ig/dL. The father did not wear personal protective equipment (PPE) 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 40-59 (ig/dL.
Of the 17 children examined, three had BLLs 5-9 (ig/dL and two had BLLs 10-19 j^ig/dL. The researchers
found many inconsistencies in adherence to PPE 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 aviation gas and found that in addition to tetraethyl Pb, the fluid was composed of various
branched and 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 4.6/106 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,
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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 jj.g/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 (Rcinard 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 (.un 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 200-220 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. 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 jj.g/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 ten metals listed as EPA air toxics (Mn, Cr, Sb, Ni, Pb, As, Co,
Cd, Se, and Be) were generally enriched in PMo.i, 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 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
Turqut et al. (2020)
Aircraft Exhaust
Cr < Zn < Ba < As < Sm < Se < V < Mg
[1.3-12.3] x 103 ng/m3
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
Cr
1,500 mg/kg
Machemer (2004)
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Source
Co-contaminants
Concentration
Reference
Iron and Steel
Manufacturing
Mn
Zn
18,000 mg/kg
5,500 mg/kg
Wildfire Ash
Co
Cu
Ni
Zn
3-11 pg/g
15-69 pg/g
6-15 pg/g
65-500 pg/g
Odiqie and Fleqal (2014)
Wildfire
Emissions
dibenzo[e,l]pyrene, 6-methylchrysene,
7,12-dimethylbenz[a]anthracene,
anthanthrene
5-methylchrysene, benzo[a]chrysene
Concentrations not
Reported
Ghetu et al. (2022)
naphtho[2,3-x]pyrene, naphtho[1,2-
b]fluoranthene
coronene, perylene, dibenzo[a,l]pyrene
House Dust Zn 876 ± 627 |jg/g Zota et al. (2011)
Near Chat Piles
Cd 4.3 ± 6.8 |jg/g
As 6.3 ± 9.9 |jg/g
Mn 143 ± 98 |jg/g
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 6 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
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 j^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%).
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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, NHANES 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
"ST
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
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. (2017).
2.1.5 Exposure Disparities for Specific Populations
1 The 2013 Pb ISA (U.S. EPA. 2013) noted elevated or differential Pb exposure and biomarker
2 levels (such as blood Pb) have been shown to be statistically related to several population characteristics,
3 including age, sex, race and ethnicity, socioeconomic status (SES), proximity to Pb sources, and
4 residential factors. The 2013 Pb ISA (U.S. EPA. 2013) evaluated past research on these population
5 characteristics' relationship to Pb exposure and biomarker levels, including biological or intrinsic (e.g.,
6 age, sex) and nonbiological or extrinsic (e.g., SES) factors. Evidence for increased exposure in this
7 section primarily relies on studies that measured BLLs. BLLs and other biomarkers are further explored
8 in Sections 2.3 and 2.4.
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2.1.5.1
Proximity to Pb Sources
The 2006 AQCD (U.S. EPA. 2006a) 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 EPA National
Air Toxics Assessment (NATA) for 2005 and industrial Pb releases obtained from the 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, 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/m3. 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 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
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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 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.
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
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percentage of non-Hispanic black, having a poverty-to-income ratio (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; Theppeang 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-
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
reported BLLs from refugees originating from many countries, whereas Ritchev etal. (2011) reported
BLLs from only Burmese refugees and Seifu 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
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1 <19 years; reference levels at which BLLs were considered elevated (either 5 or 10 (ig/dL, although some
2 studies reported both); and sampling method (i.e., venous versus capillary). There is evidence that
3 capillary blood samples may be at a higher risk of contamination and biased higher compared with venous
4 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.
(Eisenbera 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)°
<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.
(Lupone et al.,
2020)c
<16
705
2012-
2017
5 [jg/dL
10 [jg/dL
17%
Venous only.
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Study
Age
n
Study
Yr
EBLL
Reference
Value
Initial
EBLL
(%)
Follow-up
EBLL (%)
Other
EBLL
(%)*
Sampling Method
(Seaqle et al..
2020)°
<16
1,178
2010-
2015
5 |jg/dL
10 |jg/dL
8%
0.8%
Not reported.
(Seifu et al..
2020)d
<16
301
2003-
2016
5 |jg/dL
10 |jg/dL
41.9%
7.9%
Venous only.
EBLL = elevated blood lead level
""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 afterwards.
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 from Africa was 3.8 times that of children from Europe/Central Asia (reference group) and 5.6
times that of West African children specifically. 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,
and Shakva and Bhatta (2019) also reported a decrease in BLL with increased time since arrival.
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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 j^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 j^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 etal. (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 etal. (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). The 2006 Pb AQCD (U.S. EPA. 2006a) and the 2013
Pb ISA (U.S. EPA. 2013) found higher blood and bone Pb levels among African Americans.
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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. Cassidv-Bushrow
et al. (2017) measured Pb levels in tooth-matrix biomarkers among 71 children bom 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. Asians were the racial/ethnic group with the highest BLLs from 2011 to 2018.
Jones et al. (2009) found, using 1988-1991 and 1999-2004 NHANES data that although the differences
of percentage BLLs >2.5 (ig/dL between racial/ethnic groups dropped between the two sets of years, non-
Hispanic black children still had higher percentages with 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.
More recent data have shown that although average BLLs for children of all racial/ethnic groups have
continued to drop, there still exists a gap between non-Hispanic white children and minority children.
Teve 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, this gap 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. The authors also show that this gap has
decreased overtime.
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;
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
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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 j^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. 2006a) 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 etal. (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
cycle included in the study (1999-2004), although the percentage of 1- to 5-year-old children having
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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 2005-2015. 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 overtime, 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. 2006a).
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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. 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. For a detailed discussion of factors affecting particle deposition and retention in the human
respiratory tract, the reader is referred to Chapter 4 of 2019 PM ISA (U.S. EPA. 2019c). 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 equally
and totally, regardless of chemical form (Morrow et al.. 1980; Chamberlain et al.. 1978; Rabinowitz et al..
1977). Absorption half-times have been estimated for radon decay progeny in adults who inhaled aerosols
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of Pb and bismuth isotopes generated from decay of 220Rn 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 having diameters of approximately 0.3-3 nm
(Butte rweck 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.
2.2.1.1.2 Occupational Exposures
In consideration of occupational exposures, 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
cleared from the respiratory tract to the GI tract, ICRP (2017) recommends absorption fractions of 0.2
(radon and Type F) and 0.002 (Type S).
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 0.15. 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
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of the DoD modeling, NASEM (2020) supported an absorbed fraction (AF) of 0.3 from inhaled Pb as a
health-protective estimate.
Occupational studies show AFs 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 (.un 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 suggests absorption
fractions should be far greater. Using personal particle samples (open-face sampler) collected on battery
workers, Dartev etal. (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). 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
absorption fraction of 0.8 (i.e., 0.9x(0.971+0.026) + 0.15x0.006) of Pb from inhaled particles is predicted.
In this case of battery workers, assuming an absorption fraction of 0.3 could underestimate Pb absorption
by 2.5 times. Thus, it is important to consider absorption in the lung and GI tract to the extent possible to
assure accurate assessment of the risk to the exposed.
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
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
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activity, and acidic (pH 4.5) conditions. Wiseman and Zereini (2014) collected PMio, 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 (.un) 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 |am 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 (im; 0.20, 0.2 (im; 0.28, 0.06 (im).
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
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 PM10, 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 PMo 5
samples collected from areas of smelting and mining in Oruro, Bolivia. PMo 5 samples represented 79%
and 71% of PM2 5 mass in the smelting and mining area samples, respectively. The bioaccessibility
fractions of PMo 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
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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
Pb AQCD (U.S. EPA. 2006b). a single exposure to vapors of radioactive (203Pb) 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 203Pb tetramethyl Pb, 51% of the inhaled
203Pb 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 203Pb fuel [likely to have been a
mixture dominated by inorganic Pb halides but may also have included alkyl Pb species (U.S. EPA.
2006a)] are deposited and then retained in adult lungs with a half-life of 6 hours. Fifty percent of that
203Pb 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.
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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 etal.. 1998; Watson etal.. 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
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 (Ziegleretal.. 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 etal.. 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).
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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 etal.. 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
blood Pb was 0.79 (ig/dL (interquartile range: 0.51, 1.14). In a Norwegian study (Meltzer 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 (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..
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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; Juhasz et al.. 2009; U.S. EPA. 2007a; 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 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
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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
Low (<10% RBA)
Anglesite (Pb sulfate), Fe/Pb sulfate, Galena (Pb sulfide), Pb-
related sulfosalts
Medium (20-55% RBA)
Fe/Pb oxide, Fe/Pb silicate, Mine tailings, Dust and soil (mining
associated), PbO, Pb phosphate, Slag, Zn/Pb silicate
Elevated (55qn°/ rrai Pk ammuniti°n (Pb shot), Cerussite (Pb carbonate), MnPb
^ 0 ' 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.. 2006).
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 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
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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 etal.. 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
assay) of 0.75 (Drexler 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.
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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 (Hcalv ct 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 ^m
(Juhasz et al.. 2011; Kim etal.. 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 (.un. 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
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. Karnaet al. (2017) specifically examined the effect of sieving <150 |_im 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 |_im 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 etal. (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 |_im) with
the largest changes being about a 25% increase (e.g., from 38 to 63%) going from <250 to <50 |_im
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1 particles. However, the largest bioaccessibility change was in the opposite direction from soil collected at
2 a shooting range, where bioaccessibility increased from about 69 to 99%, going from <50 to <250 (.un
3 particles. Overall, among studies, there are no consistent changes in IVBA as a function of particle size.
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
4
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
5 A simple conceptual representation of Pb distribution is that it contains a fast turnover pool,
6 comprising mainly soft tissue, and a slow pool, comprising mainly skeletal tissues (Rabinowitz et al..
7 1976). The highest soft tissue concentrations in adults occur in liver and kidney cortex (Gerhardsson et
8 al.. 1995; Oldereid et al.. 1993; Gerhardsson et al.. 1986; Barry. 1975; Gross et al.. 1975). Pb in blood
9 (i.e., plasma) exchanges with both of these compartments.
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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 etal.. 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
molecular weight of <10 kDa (Bergdahl etal.. 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 j^ig/dL RBCs (or -40 (ig/dL whole
blood), and the apparent dissociation constant has been estimated at -1.5 j^ig/L (Bergdahl etal.. 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 nonoccupational^ 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). Sobinetal. (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 |ag/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.
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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;
Bcrgdahl 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
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
O)
Ł i.o
CO
E
V)
ro
o- 0.5
0.0
0
oAdults •Children
o ** •
°o Jp^bo °#
20 40 60 80
Blood Pb (|jg/dL)
100
Source: Adapted with permission of Elsevier Publishing and the Finland Institute of Occupational Health, Bergdahl et al. (1999);
Bergdahl et al. (1997b).
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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
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 (1986).
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 Leggett. 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
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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 (Bcrgdahl 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
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.
0.60
0.50 -
1) 0.40 -
n
0.30 -
ra
E
___
ra 0.20 A
0.10 -I
0.00
10 20 30 40 50
Blood Pb (iig/dL)
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50 -
5
O)
.Q
CL
"D
O
O
40 -
30 -
CQ 20 -
10 -
100 200 300
Intake (tig/day)
1.0
0.8
0.6
5
Si
v>
3
0)
¦0
O"
O-* Ł
0.2
0.0
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 (Leggett. 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.
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
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the bone [see 1(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
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. 2006a) 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'Flahertv. 1995). Thus, most (60-65%) of the
Pb acquired early in life is not permanently fixed in the bone (O'Flahertv. 1995; Lcggctt. 1993; ICRP.
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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\x 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
inducer of the protein in comparison with the inducers Cd and Zn (Waalkes 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 (Lcggctt. 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; Azaretal.. 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 etal.. 2009; Patel and Prabhu. 2009;
Carbone etal.. 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
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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 (Harvillc 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 204Pb 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
pregnancy (1,500 |_ig Pb/kg/day; 7 days/week), the range was lower (7-25%) (O'Flahertv. 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. 2006b'). 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 203Pb
tetraethyl or tetramethyl Pb (1 mg/m3), -50% of the 203Pb body burden was associated with liver and 5%
with kidney; the remaining 203Pb was widely distributed throughout the body (Heard et al.. 1979). The
kinetics of 203Pb 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
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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 (Lcggctt. 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
(Araki et al.. 1986) and plasma Pb concentration (Rentschler et al.. 2012; Bergdahl etal.. 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; Vanderetal.. 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; Berger et al.. 1990; Kosteretal.. 1989; Manton and Mallov. 1983; Ryu et al.. 1983;
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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 2009-2016 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
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 etal.. 1969; Hursh et al.. 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 tm =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 et al.. 1979). Fecal:urinary excretion ratios were 1.8 following exposure to
tetraethyl Pb and 1.0 following exposure to tetramethyl Pb (Heard etal.. 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).
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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
methods because of its ease of use. K-shell XRF (KXRF) 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.. 2011). Zhang et al. (2021)
evaluated a portable XRF device against a traditional KXRF instrument using the mid-tibia bone in 71
people of three Indiana communities. The correlation between the portable XRF and KXRF 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.
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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 overthe 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
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.
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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. 2006a) 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
additional summaries of what is known regarding the use of teeth, hair, saliva, and serum 8-
aminolevulinic acid (8-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 ((.i-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 (Rabinowitz et al.. 1993). The neonatal line
formed in deciduous teeth during birth can be used to distinguish between prenatal and postnatal dentine
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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.
Aroraetal. (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 (statistically
significant) and postnatal teeth (p = 0.056). Wang et al. (2017b) applied (i-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. 2006a) 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
predictor of blood Pb (U.S. EPA. 2013). Pb concentrations have been found to vary along the hair shaft
(Jursaet 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).
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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 j^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.
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 etal.. 1996; Gulsonetal.. 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 tin = 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
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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 having 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
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
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ceased (Fleming et al.. 1997; Inskip et al.. 1996; Smith etal.. 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
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 (|_ig 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
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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-time1 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
(Spccht 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-shell XRF tibia bone Pb was well
correlated with blood Pb (r2 = 0.59; n = 157). The correlation between K-shell 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. 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).2 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.
1 Half-time (months) = [Ln( 1/2)x(Time to 10 |ig/dL)| / Ln[(10 |ig/dL)/(midpoint blood Pb)]
2 These 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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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 of Specht et al. (2019b).
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 reduced their
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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 tm 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 etal. (2011)
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,
blood and bone Pb may show relatively strong associations. A recent analysis of a subset of data from the
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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).
25 30 35 40 45 50 55 60 65 70
Age (year)
3
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Age (year)
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Note: A constant baseline intake of 20 |jg/day from age 0-30 results in a quasi-steady state blood Pb concentration and body
burden. An increase in 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. 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 (Leqqett. 1993).
Figure 2-8 Simulation of relationship between blood Pb concentration, bone
Pb, and body burden in adults.
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; Gulson et
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; Gulsonetal.. 1997; Lagerkvist et al.. 1996; Schuhmacher et al..
1996; Rothcnbcrg 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 (Ettingcr 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 j^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
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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,
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 jj.g/g
bone and is nearly double the rate found in premenopausal women (0.07-0.08 (ig/dL per jj.g/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).
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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 etal.. 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).
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
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1 described previously for blood Pb, the relationships between urinary Pb and bone Pb will diverge with
2 increasing exposure duration and age, even if exposure remains constant. Furthermore, following an
3 abrupt cessation of exposure, urine Pb will quickly decrease while bone Pb will remain elevated (Figure
4 2-11, lower panel).
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w
o
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0 2 4 6 8 10
Age (year)
Blood
Brain
40 45 50 55 60
Age (year)
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 (Leggett. 1993).
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
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ra 3.
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Blood
Urine
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r
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*
*
*
/
/
*
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1 1 1 1
1 1 1
20
-- 15
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25 30 35 40 45 50 55 60 65 70
Age (year)
Note: For the upper panel, model simulations are for a 40-year-old having a constant intake from birth of between 1 and
1,000 |jg/day. For the lower panel, the 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
(Leqqett. 1993).
Figure 2-11 Relationship between Pb in urine, plasma, blood, and bone.
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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 (|jg/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 (|jg/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
Males
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
Females
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 Americans
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 blacks
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 whites
2011-2012
0.993
0.914
, 1.08
2,493
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Survey Stratum
Period
Geometric Mean (|jg/dL)
95% Confidence Interval
Number of Subjects
2013-2014
0.882
0.820, 0.950
1,848
2015-2016
0.835
0.774, 0.900
1,511
2017-2018
0.772
0.731, 0.816
2,536
All Hispanics
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
Asians
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 = years
Limit 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
jŁ 1.75
"D
CO
0)
^ 1.25
O
O
CO
0.75
T
* 9
• Children 1-5 yrs
A Children 6-11 yrs
~ Teens 12-19 yrs
¦ Adults > 20 yrs
0.25
99-00 01-02 03-04 05-06
07-08 09-10
Survey Period
11-12 13-14 15-16 17-18
Note: Shown are geometric means and 95% CIs based on data from NHANES IV CDC (2021a).
Figure 2-12 Temporal trend in blood Pb concentrations.
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COHORT BIRTH YEARS
1900-
1930-
1940-
1950-
1960-
1970-
1975-
1980-
1985-
1990-
1995-
2000-
2005-
2010-
2015-
1929
1939
¦1949
1959
1969
¦1974
1979
1984
1989
1994
1999
2004
2009
2014
2019
\°P \°P ^ ^ ^ ^
EXAM YEAR
Note: The means of logged blood Pb were weighted to represent national averages. Data were from the publicly available NHANES
II. 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.
Figure 2-13 Blood Pb cohort means versus year of exam.
3.0
5 2.5
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Figure 2-14 Blood Pb geometric means versus year of NHANES exam by
race/ethnicity.
Additional analyses have used NHANES data to investigate the decline in BLLs overtime. 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 a blood 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. 2006a. 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, aviation gas, 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://cfpub.epa.gov/ncea/isa/recordisplav.cfm?deid=357282). 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.. 2017: 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,
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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
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 (RCTs) and quasi-RCTs (Nussbaumcr-Strcit et al.. 2020; Nussbaumer-Streit et al.. 2016;
Yeoh etal.. 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
B I 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 jjg/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, NE.
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 (.ig/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
(https://cfbub.epa.gov/ncea/isa/recordisplav.cfm'?deid=357282), aviation fuel remains a major source of
Pb emissions in ambient air. Miranda et al. (2011) found a monotonically decreasing trend in BLLs of
children aged 9 months to 7 years and distance from airports in six counties m 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. Zahrari
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
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temporally to 448 airports in Michigan and emission inventories of Pb-releasing industry sites found in
the TRI. 40.2% of measurements were at/below detection levels. 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 j^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 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 (NASCAR) and the Automobile Racing Club of America (ARCA) 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
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effects of living in a race 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 EPA AQS at monitors within 4 km of NHANES participants. The authors
generally found positive associations between BLLs for all five age groups (from 1-5 years to >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 (Miclkc et al.. 2019b; Mielke et
al.. 2019a. 2017; Rabito et al.. 2012; Mielke et al.. 2011b; Zahranetal.. 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 amonotonic
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 and year-long tap water filter samples from the interiors of seven
homes. 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 (ig/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. Spanieretal. (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 jj.g/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 Gracf. 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 of April 25, 2014 to 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 j^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 after the
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 etal. (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 jj.g/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 jxg/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.
Wilker etal. (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 KXRF 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 jj.g/g (SD = 19.9) and a mean tibia Pb measurement of
21.6 jj.g/g (SD = 13.6) at the 1 st 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 overtime, 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 AQCD (U.S. EPA. 2006a). 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
(M9 Pb/g CRf
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
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Survey Stratum
Period
Geometric Mean
(|jg Pb/g CRf
95% Confidence Interval
Number of
Subjects
12-19 yr
2011-2012
0.259
0.219, 0.305
390
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
blacks
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
whites
2013-2014
0.267
0.245, 0.290
985
2015-2016
0.275
0.247, 0.305
924
All Hispanics
2011-2012
0.372
0.327, 0.423
573
2013-2014
0.270
0.239, 0.305
701
2015-2016
0.284
0.258, 0.312
982
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O O* * n . . Geometric Mean Number of
Survey Stratum Period 95% Confidence Interval „ . . .
(|jg Pb/g CR)a Subjects
Asians 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 = years
aValues are in |jg Pb/g creatine (CR).
Source: Data sourced from CDC (2021a)
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 208Pb:43Ca, were 4.104/ 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 jj.g/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.
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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) 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 the rate of 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 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 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
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(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.
2.5.1 Air Pb-Blood Pb Relationships in Children
Within the literature and 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/ni3 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/m3. 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 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), with recognition of uncertainty and variability in the air-to-blood
relationship, air-to-blood slopes of 5, 7, and 10 j^ig/dL per (.ig/ni3 were used in evaluating air-related IQ
loss of children (see Table 4 of that rule). The highest slope of 1:10 (ig/dL per (ig/m3 was used to reflect
the lower BLLs of children and PbA more recently observed in the United States relative to when most of
the studies describing the air-to-blood relationship had been conducted.
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/m3. The air Pb-blood Pb relationship in terms of slope factors or ratios was not discussed in the
2006 Pb AQCD (U.S. EPA. 2006a). 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/m3. 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
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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 a biokinetic 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; Haves 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).
An inherent assumption in univariate regression models relating blood Pb to air Pb is that each population
has the same relative contribution to blood Pb from other (non-air) pathways or that the non-air
contribution is negligible relative to air-related pathways. For the older studies, there may be a greater
likelihood of representing the effect of air Pb on blood Pb due to relatively less interference from non-air
exposure pathways. However, overarching distinctions between old and new studies should be made with
caution given that Pb in other media has also decreased over time (e.g., due to banning the use of Pb
solder in food cans and plumbing).
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
compare two populations for which all Pb sources but PbA concentrations are unchanged. Such a nearly
ideal study, Hilts (2003) reported the change in PbB from 1996 to 2001 for children under 5 yrs 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 understanding the PbB-PbA association came from leveraging the USEPA AQS with
NHANES surveys. The PbB-PbA 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 is a representative sample of the U.S. population. However,
merging PbB results from multiple NHANES periods with the USEPA 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.
x x [PM]*""')
Equation 2-2
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where: b is the logarithm base (either 2.7183 or 10) used in a study, Po is the regression intercept, PpbA 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/nr1 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 (Mcng 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 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).
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 |jg/dL: 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
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Reference
Study Methods
Model Description
Blood Pb-Air Pb Slope3
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;
Meng et al. n = 2,150, PM10), 6-11 yr (n = 212,
(2014) TSP; n = 2,261, PM10)
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/m3 and
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)9
37.7 (PM10, 0.0054)
6-11 yr
3.0 (TSP, 0.0151)
20.1 (PM10, 0.0051)
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)'
8.6 (0.75)'
Tripathi et al.
(2001)
Location: Mumbai, India (multiple
residential locations)
Yr: 1984-1996
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)
3.6 (0.45)k
Richmond-
Bryant 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)
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Reference
Study Methods
Model Description
Blood Pb-Air Pb Slope3
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°
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
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Reference
Study Methods
Model Description
Blood Pb-Air Pb Slope3
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 = years
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 2014); (Richmond-Bryant et al.. 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) x 0.09 + 0.58
cln(PbB) = In(PbA) x 0.3485 + 2.853
dln(PbB) = In(PbA) x 0.2159 + 2.620
eln(PbB) = In(PbA) x 0.24 + 3.17
fPbB = PbA x 7.0
91-5 years [TSP, In(PbB) = ln(Pb A) x 0.056 + 1.024; PM10, In(PbB) = ln(Pb A) x 0.104 + 1.213]; 6-11 years [TSP, In(PbB) = ln(Pb
A) x 0.028 + 0.596; PM10, In(PbB) = ln(Pb A) x 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 = Pb A x 3.6
'1-5 years [1988-1994, In(PbB) = ln(Pb A) x 0.1395 + 1.9315; 1999-2008, In(PbB) = ln(Pb A) x 0.0755 + 1.1419]; 6-11 years
[1988-1994, In(PbB) = ln(Pb A) x 0.1535 + 1.8118; 1999-2008, In(PbB) = ln(Pb A) 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 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.
pPbB = PbA x 9.63
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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, PMio, and PM2 5 from the 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 EPA study (Cho etal.. 2011) showing the mass median diameter of airborne Pb had shifted
from <2.5 |_im priorto the phase-out of leaded gasoline to somewhere between 2.5 and 10 |_im 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. The
data for PM2 5 are not presented in Table 2-13 or Figure 2-16 since there was no significant relationship
between blood Pb and air Pb for any age group. The lack of a significant relationship for PM2 5 may be
attributed to airborne Pb being found associated with particles larger than 2.5 |_im (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%). To derive slope factors at the central
tendency for air Pb, it was necessary to solve for Po (see Equation 2-2) using data from Tables 2 and 3 of
Meng et al. (2014). It was also necessary to use the GM air Pb from Table 1 of Meng et 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 PM10 than
either TSP or PM2 5.
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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 EPA AQS Pb-PMio data because their
prior work (Mcng 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 and elements used to calculate atmospheric soil concentrations were obtained from four TSP
sampling sites in 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/nr1 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 (Ngucta et al.. 2014; Deshommes et al.. 2013). Although magnitude of water Pb concentrations
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 Ngucta 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
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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 of Zahran ct 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);
r2 = 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/m3. 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 PpbAwas 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 ((ig/dL per (ig/m3) estimated for this study by rank are
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.
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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. 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 amaximum daily value of 210 (.ig/m3. Hand-wipe samples were collected and
analyzed for Pb (GM = 793 jj.g, 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/m3 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/m3 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
crystal workers, it averaged 21.9 (ig/dL. Both studies employed log-log regression models, resulting in
slopes of 0.49 (Pierre et al.. 2002) and 0.08 (Lai et al.. 1997).
2.5.3 Soil Pb-Blood Pb Relationships
Slope factor models represent the 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)
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provides a description of these models and past reviews of environmental Pb-blood Pb data to develop
these relationships. The EPA Adult Lead Methodology is a slope factor model that has had extensive use
in the EPA Superfiind Program. The 2006 AQCD (U.S. EPA. 2006a) 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 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.
A
Blood Pb 2000-2005 vs.
¦
Soil Pb 1998-2001
A .
a All
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(WJT 0 2^ % o Om 0 °
Blood Pb 2011-2016 vs. Soil Pb 2013-2017
^ 5 » "iT"" o
• Flooded
oUnflooded
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y =
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y = 0.47ln(x) - 0.31
1 1 1
R2 = 0.581
R2 = 0.488
1 1 1
200
400
600
800 1000 1200
Median Soil Pb (mg/kg)
1400
1600
1800
2000
Figure 2-17 Blood Pb versus soil Pb for two New Orleans surveys completed
in 2001 and 2017.
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1.0 i
CD
Q_
_0 . .
uo W)
_Q
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~o E
s
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•- Mielke et al. 2019 (old data}
Zahran et al. 2011
_Q
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Soil Lead (mg/kg)
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=-i
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1.00 T
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exceed CDC's BLRV or, formerly, the "level of concern") identified using capillary 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. (202 la). Egendorf et al. (202 la) 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,
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including those that may be outside of the domain of paired predictor-outcome data used to develop the
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) consists of a systemic biokinetics
model (Lcggctt. 1993) and absorption factors for inhaled Pb (ICRP. 1995). The Leggett model 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 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.
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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 AQCD (U.S. EPA. 2006a) provides
an introduction to the IEUBK model along with components. 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 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. 2006a; Maddaloni et al.. 2005; O'Flahertv et al.. 1998; O'Flahertv. 1998; Pounds and Leggett.
1998; Leggett. 1993; Leggett et al.. 1993). Section 4.4.8 of the 2006 AQCD (U.S. EPA. 2006a) provides
an introduction to 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 an 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).
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,
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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. 2006a). 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
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.
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There is evidence Pb concentrations in the environment have decreased within the United States
over time. As noted in Appendix 1, air Pb concentration in the United States has 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 make
a contribution 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 (.un 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, 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
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
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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.
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
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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
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).
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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 (Mcng 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.
Although slope factors increase with decreasing air Pb concentration, it is possible that
contribution from non-air exposure pathways may lead to the high slope factors at low air concentrations.
In other words, in older studies in which leaded gasoline or local sources were a major contributor to air
Pb, there may be a greater likelihood of discerning the true effect of air Pb on blood Pb due to relatively
less contribution from non-air exposure pathways. However, overarching distinctions between old and
new studies should be made with caution given that Pb in all media, not just air, has decreased over time.
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